Sediment Dredging at Superfund Megasites: Assessing the - CLU-IN

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Committee on Sediment Dredging at Superfund Megasites
Board on Environmental Studies and Toxicology
Division on Earth and Life Studies
THE NATIONAL ACADEMIES PRESS
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NOTICE: The project that is the subject of this report was approved by the Governing
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of Medicine. The members of the committee responsible for the report were chosen for
their special competences and with regard for appropriate balance.
This project was supported by Contract 68-C-03-081 between the National Academy of
Sciences and the U.S. Environmental Protection Agency. Any opinions, findings, conclusions, or recommendations expressed in this publication are those of the authors and do
not necessarily reflect the views of the organizations or agencies that provided support for
this project.
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Photograph of environmental dredging to remove contaminated sediments in the Fox
River, Wisconsin. Courtesy of Gregory A. Hill, Wisconsin Department of Natural Resources.
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www.national-academies.org
COMMITTEE ON SEDIMENT DREDGING AT SUPERFUND MEGASITES
Members
CHARLES R. O’MELIA (Chair), Johns Hopkins University, Baltimore, MD
G. ALLEN BURTON, Wright State University, Dayton, OH
WILLIAM H. CLEMENTS, Colorado State University, Fort Collins
FRANK C. CURRIERO, Johns Hopkins Bloomberg School of Public Health, Baltimore, MD
DOMINIC DI TORO, University of Delaware, Newark
NORMAN R. FRANCINGUES, OA Systems Corporation, Vicksburg, MS
RICHARD G. LUTHY, Stanford University, Stanford, CA
PERRY L. MCCARTY, Stanford University, Stanford, CA
NANCY MUSGROVE, Management of Environmental Resources, Inc., Seattle, WA
KATHERINE N. PROBST, Resources for the Future, Washington, DC
DANNY D. REIBLE, University of Texas, Austin
LOUIS J. THIBODEAUX, Louisiana State University, Baton Rouge
DONNA J. VORHEES, The Science Collaborative, Ipswich, MA
JOHN R. WOLFE, Limno-Tech, Inc., Ann Arbor, MI
Project Staff
KARL E. GUSTAVSON, Project Director
RAYMOND A. WASSEL, Program Director
NORMAN GROSSBLATT, Senior Editor
MIRSADA KARALIC-LONCAREVIC, Manager, Technical Information Center
MORGAN MOTTO, Senior Project Assistant
RADIAH ROSE, Senior Editorial Assistant
Sponsor
U.S. ENVIRONMENTAL PROTECTION AGENCY
v
BOARD ON ENVIRONMENTAL STUDIES AND TOXICOLOGY1
Members
JONATHAN M. SAMET (Chair), Johns Hopkins University, Baltimore, MD
RAMÓN ALVAREZ, Environmental Defense, Austin, TX
JOHN M. BALBUS, Environmental Defense, Washington, DC
DALLAS BURTRAW, Resources for the Future, Washington, DC
JAMES S. BUS, Dow Chemical Company, Midland, MI
RUTH DEFRIES, University of Maryland, College Park
COSTEL D. DENSON, University of Delaware, Newark
E. DONALD ELLIOTT, Willkie Farr & Gallagher LLP, Washington, DC
MARY R. ENGLISH, University of Tennessee, Knoxville
J. PAUL GILMAN, Oak Ridge Center for Advanced Studies, Oak Ridge, TN
SHERRI W. GOODMAN, Center for Naval Analyses, Alexandria, VA
JUDITH A. GRAHAM, American Chemistry Council, Arlington, VA
WILLIAM P. HORN, Birch, Horton, Bittner and Cherot, Washington, DC
WILLIAM M. LEWIS, JR., University of Colorado, Boulder
JUDITH L. MEYER, University of Georgia, Athens
DENNIS D. MURPHY, University of Nevada, Reno
PATRICK Y. O’BRIEN, ChevronTexaco Energy Technology Company, Richmond, CA
DOROTHY E. PATTON (retired), Chicago, IL
DANNY D. REIBLE, University of Texas, Austin
JOSEPH V. RODRICKS, ENVIRON International Corporation, Arlington, VA
ARMISTEAD G. RUSSELL, Georgia Institute of Technology, Atlanta
ROBERT F. SAWYER, University of California, Berkeley
KIMBERLY M. THOMPSON, Massachusetts Institute of Technology, Cambridge
MONICA G. TURNER, University of Wisconsin, Madison
MARK J. UTELL, University of Rochester Medical Center, Rochester, NY
CHRIS G. WHIPPLE, ENVIRON International Corporation, Emeryville, CA
LAUREN ZEISE, California Environmental Protection Agency, Oakland
Senior Staff
JAMES J. REISA, Director
DAVID J. POLICANSKY, Scholar
RAYMOND A. WASSEL, Senior Program Officer for Environmental Sciences and Engineering
KULBIR BAKSHI, Senior Program Officer for Toxicology
EILEEN N. ABT, Senior Program Officer for Risk Analysis
KARL E. GUSTAVSON, Senior Program Officer
ELLEN K. MANTUS, Senior Program Officer
SUSAN N.J. MARTEL, Senior Program Officer
STEVEN K. GIBB, Program Officer for Strategic Communications
RUTH E. CROSSGROVE, Senior Editor
1This study was planned, overseen, and supported by the Board on Environmental
Studies and Toxicology.
vi
OTHER REPORTS OF THE
BOARD ON ENVIRONMENTAL STUDIES AND TOXICOLOGY
Models in Environmental Regulatory Decision Making (2007)
Toxicity Testing in the Twenty-first Century: A Vision and a Strategy (2007)
Environmental Impacts of Wind-Energy Projects (2007)
Scientific Review of the Proposed Risk Assessment Bulletin from the Office of
Management and Budget (2007)
Assessing the Human Health Risks of Trichloroethylene: Key Scientific Issues (2006)
New Source Review for Stationary Sources of Air Pollution (2006)
Human Biomonitoring for Environmental Chemicals (2006)
Health Risks from Dioxin and Related Compounds: Evaluation of the EPA Reassessment
(2006)
Fluoride in Drinking Water: A Scientific Review of EPA’s Standards (2006)
State and Federal Standards for Mobile-Source Emissions (2006)
Superfund and Mining Megasites―Lessons from the Coeur d’Alene River Basin (2005)
Health Implications of Perchlorate Ingestion (2005)
Air Quality Management in the United States (2004)
Endangered and Threatened Species of the Platte River (2004)
Atlantic Salmon in Maine (2004)
Endangered and Threatened Fishes in the Klamath River Basin (2004)
Cumulative Environmental Effects of Alaska North Slope Oil and Gas Development (2003)
Estimating the Public Health Benefits of Proposed Air Pollution Regulations (2002)
Biosolids Applied to Land: Advancing Standards and Practices (2002)
The Airliner Cabin Environment and Health of Passengers and Crew (2002)
Arsenic in Drinking Water: 2001 Update (2001)
Evaluating Vehicle Emissions Inspection and Maintenance Programs (2001)
Compensating for Wetland Losses Under the Clean Water Act (2001)
A Risk-Management Strategy for PCB-Contaminated Sediments (2001)
Acute Exposure Guideline Levels for Selected Airborne Chemicals (five volumes,
2000-2007)
Toxicological Effects of Methylmercury (2000)
Strengthening Science at the U.S. Environmental Protection Agency (2000)
Scientific Frontiers in Developmental Toxicology and Risk Assessment (2000)
Ecological Indicators for the Nation (2000)
Waste Incineration and Public Health (2000)
Hormonally Active Agents in the Environment (1999)
Research Priorities for Airborne Particulate Matter (four volumes, 1998-2004)
The National Research Council’s Committee on Toxicology: The First 50 Years (1997)
Carcinogens and Anticarcinogens in the Human Diet (1996)
Upstream: Salmon and Society in the Pacific Northwest (1996)
Science and the Endangered Species Act (1995)
Wetlands: Characteristics and Boundaries (1995)
Biologic Markers (five volumes, 1989-1995)
Review of EPA's Environmental Monitoring and Assessment Program (three volumes,
1994-1995)
Science and Judgment in Risk Assessment (1994)
vii
Pesticides in the Diets of Infants and Children (1993)
Dolphins and the Tuna Industry (1992)
Science and the National Parks (1992)
Human Exposure Assessment for Airborne Pollutants (1991)
Rethinking the Ozone Problem in Urban and Regional Air Pollution (1991)
Decline of the Sea Turtles (1990)
Copies of these reports may be ordered from the National Academies Press
(800) 624-6242 or (202) 334-3313
www.nap.edu
viii
Preface
Contaminated sediments in aquatic environments pose health risks to humans and
other organisms. Nationwide, the full extent of the problem is poorly documented, but it is
well known that rivers, harbors, lakes, and estuaries fed by current or former industrial,
agricultural, or mining areas frequently contain contaminated sediments. It is also well
known that contaminants in the sediments can directly harm aquatic organisms or accumulate in their tissues, which can be consumed by humans. The potential adverse effects on
human health and the environment are compelling reasons to reduce exposures.
From a regulatory standpoint, contaminated sediments are challenging to manage.
The Superfund program, administered by the U.S. Environmental Protection Agency
(EPA), is intended to protect human health and the environment from sites contaminated
with hazardous substances. An array of techniques are available for remediating contaminated sediments, each with advantages and disadvantages. Decisions about which remedial measures to implement and, in particular, whether to dredge at contaminated sediment sites have proved to be among the most controversial at Superfund megasites. The
scientific and technical difficulties of deciding on a remedial option are augmented by the
challenges of implementing a regulatory authority that holds responsible parties liable for
paying for the cleanup.
Regardless of cost or controversy, achieving the expected effect of remedial actions—improvements in the environment—is of primary importance. That is true for regulators who may require cleanup of a site, parties responsible for funding the cleanup, and
communities and user groups affected by the contamination. This report, one piece of a
continuing dialogue, seeks to assess the effectiveness of environmental dredging in reducing risks associated with contaminated sediments, particularly at large, complex Superfund
sites (these sites are termed “megasites” when the cost of remedial activities is anticipated
to exceed $50 million).
Over the course of its study, the Committee on Sediment Dredging at Superfund
Megasites held three public sessions at which it heard presentations on dredging projects
and received input from members of the public and other interested parties; two closed,
deliberative sessions were held over the course of the year-long study. The report consists
of six chapters. Chapters 1 through 3 introduce the problem, provide background on the
issues, and describe the committee’s approach to addressing the statement of task. Chapter
4 considers the data on various dredging sites to develop recommendations for implement-
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x
Preface
ing sediment-management techniques. Chapter 5 evaluates current monitoring approaches
and suggests future approaches. Finally, Chapter 6 takes a broader look and considers
sediment management at the national level, and it provides conclusions and recommendations to improve decision-making in the future.
The committee gratefully acknowledges the following for making presentations and
for providing information during the committee’s meetings: Loretta Beaumont, U.S. House
of Representatives; Stephen Ells, Leah Evison, Elizabeth Southerland, Dave Dickerson,
James Brown, Young Chang, William “Skip” Nelson, and Marc Greenberg, U.S. Environmental Protection Agency; Michael Palermo, U.S. Army Corps of Engineers (retired); Clay
Patmont, Anchor Environmental; Steven Nadeau, Sediment Management Work Group;
John Connolly, QEA; Larry McShea, Alcoa; Rick Fox, Natural Resource Technology; Mike
Jury, CH2M Hill; John Kern, KERN Statistical Services; and Todd Bridges, U.S. Army
Corps of Engineers.
The committee is also grateful for the assistance of National Research Council staff
in preparing this report: Karl Gustavson, study director; James Reisa, director of the Board
on Environmental Studies and Toxicology; Ray Wassel, program director; Norman Grossblatt and Ruth Crossgrove, senior editors; Mirsada Karalic-Loncarevic, manager of the
Technical Information Center; Morgan Motto, senior project assistant; and Radiah Rose,
senior editorial assistant. Finally, I thank the members of the committee for their dedicated
efforts throughout the development of this report.
Charles R. O’Melia, Chair
Committee on Sediment Dredging at
Superfund Megasites
Acknowledgment of Review Participants
This report has been reviewed in draft form by persons chosen for their diverse perspectives and technical expertise in accordance with procedures approved by the National
Research Council (NRC) Report Review Committee. The purpose of this independent review is to provide candid and critical comments that will assist the institution in making its
published report as sound as possible and to ensure that the report meets institutional
standards of objectivity, evidence, and responsiveness to the study charge. The review
comments and draft manuscript remain confidential to protect the integrity of the deliberative process. We wish to thank the following for their review of this report:
Todd S. Bridges, U.S. Army Engineer Research and Development Center
Edwin H. Clark II, Earth Policy Institute
J. Paul Doody, Blasland, Bouck & Lee, Inc., an Arcadis Company
Rick Fox, Natural Resource Technology, Inc.
Paul Fuglevand, Dalton, Olmsted & Fuglevand, Inc.
John C. Henningson, Henningson Environmental Services
Chris Ingersoll, U.S. Geological Survey
Stephen U. Lester, Center for Health & Environmental Justice
Jeffrey S. Levinton, Stony Brook University
Victor S. Magar, ENVIRON
Larry McShea, Alcoa
Steven C. Nadeau, Honigman Miller Schwartz and Cohn, LLP
Clay Patmont, Anchor Environmental, LLC.
Harlee S. Strauss, H. Strauss Associates, Inc.
Although the reviewers listed above have provided many constructive comments
and suggestions, they were not asked to endorse the conclusions or recommendations, nor
did they see the final draft of the report before its release. The review of this report was
overseen by George M. Hornberger, University of Virginia, and C. Herb Ward, Rice University. Appointed by the NRC, they were responsible for making certain that an independent examination of this report was carried out in accordance with institutional procedures and that all review comments were carefully considered. Responsibility for the final
content of this report rests entirely with the author committee and the institution.
xi
Abbreviations
AOCs
ARARs
ARCS
AVS
BMPs
CFR
COCs
COPCs
CSO
DGPs
ELISA
EPA
EQC
KDGPS
MC
MCSS
MNR
MTCA
NHANES
NNS
NOAA
NPL
NSI
OMC
PAHs
PCBs
PED
PNEC
PRA
RAL
RI/FS
ROD
areas of concern
applicable or relevant and appropriate requirements
Assessment and Remediation of Contaminated Sediments
acid volatile sulfide
best management practices
Code of Federal Regulations
contaminants of concern
contaminants of potential concern
combined sewer outfalls
differential global positioning systems
enzyme-linked immunosorbent assays
U.S. Environmental Protection Agency
environmental quality criteria
kinematic differential global positioning systems
Main Channel
Major Contaminated Sediment Sites Database
monitored natural recovery
Model Toxics Control Act
National Health and Nutrition Examination Survey
Northern Near Shore
National Oceanographic and Atmospheric Administration
National Priorities List
National Sediment Inventory
Outboard Marine Corporation
polycyclic aromatic hydrocarbons
polychlorinated biphenyls
polyethylene device
probable no effect concentration
probabilistic risk assessment
remedial action level
remedial investigation and feasibility study
record of decision
xiii
xiv
ROPS
SMA
SMS
SMU
SPI
SPMD
SPME
SQG
SQO
TSCA
UCL
Abbreviations
remedial options pilot study
sediment management area
Sediment Management Standards
Sediment Management Unit
sediment profile imagery
semipermeable membrane device
solid-phase microextraction
sediment quality guidelines
sediment quality objectives
Toxic Substances Control Act
upper confidence limit
Contents
SUMMARY ..........................................................................................................................................1
1
INTRODUCTION ...................................................................................................................16
The Challenge of Contaminated Sediments, 16
The Charge to the Committee on Sediment Dredging at Superfund Megasites, 18
National Research Council and the Committee Process, 19
Report Organization, 20
References, 21
2
SEDIMENT MANAGEMENT AT SUPERFUND MEGASITES ...................................23
Overview of Superfund and Sediment Megasites, 23
Evaluating Risk Reduction at Contaminated Sediment Sites, 34
Contaminated Sediment Management Techniques, 45
Historical Perspective on the Use of Removal Technologies to Reduce Risk, 46
Overview of Environmental Dredging, 48
Technical Issues Associated with Dredging, 54
References, 63
3
EFFECTIVENESS OF ENVIRONMENTAL DREDGING IN REDUCING RISK:
FRAMEWORK FOR EVALUATION...................................................................................70
Sources of Information About Environmental-Dredging Projects, 72
Criteria Used to Select Environmental Dredging Projects for Effectiveness
Evaluation, 75
Selection of Environmental-Dredging Projects, 76
Approach to Evaluating Dredging Projects, 83
References, 86
4
EVALUATION OF DREDGING EFFECTIVENESS:
WHAT HAS EXPERIENCE TAUGHT US? ........................................................................90
Introduction, 90
Data Availability and Accessibility, 92
Dredging Effectiveness, 95
Factors Affecting Dredging Effectiveness, 105
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xvi
Contents
Management of Design and Implementation to Maximize Dredging
Effectiveness, 138
Conclusions, 162
Recommendations, 164
References, 165
5
MONITORING FOR EFFECTIVENESS: CURRENT PRACTICES AND
PROPOSED IMPROVEMENTS .........................................................................................178
Monitoring for Effectiveness, 178
Monitoring Principles, 179
Current Monitoring Practices, 184
Monitoring-Program Design, 195
Data Sufficiency and Statistical Design, 214
Apprroaches to Improving Monitoring, 217
Conclusions, 224
Recommendations, 226
References, 227
6
DREDGING AT SUPERFUND MEGASITES:
IMPROVING FUTURE DECISION MAKING ..............................................................240
Introduction, 240
Managing Sediment Megasites in the Future, 241
Conclusions and Recommendations, 248
References, 261
APPENDIXES
A
STATEMENT OF TASK FOR THE COMMITTEE ON SEDIMENT
DREDGING AT SUPERFUND MEGASITES........................................................264
B
BIOGRAPHIC INFORMATION ON THE COMMITTEE ON SEDIMENT
DREDGING AT SUPERFUND MEGASITES........................................................266
C
SUMMARY OF REMEDIAL ACTION OBJECTIVES, CLEANUP LEVELS
(NUMERICAL REMEDIAL GOALS), AND THEIR ACHIEVEMENT AT
SEDIMENT-DREDGING SITES...............................................................................274
BOXES, TABLES, AND FIGURES
BOXES
S-1
2-1
2-2
2-3
Overview of Conclusions and Recommendations, 5
Evaluation Criteria for Superfund Remedial Alternatives, 25
How Large Is a Megasite? 31
Presidential/Congressional Commission on Risk Assessment and Risk
Management, 38
Contents
2-4
2-5
2-6
2-7
2-8
2-9
3-1
3-2
3-3
3-4
4-1
4-2
4-3
4-4
4-5
4-6
5-1
5-2
5-3
5-4
5-5
5-6
5-7
5-8
5-9
6-1
6-2
6-3
xvii
Eleven Principles of Contaminated Sediment Management, 43
Remedial Approaches to Contaminated Sediment in Situ Approaches, 47
Objectives of Environmental Dredging, 49
Equipment Commonly Used in Environmental Dredging, 51
Site-Specific Factors Affecting Resuspension, Release, and Residuals Sediment
Physical and Chemical Properties, 59
Specific Processes Contributing to the Residual Layer During Dredging, 61
Framework for Evaluating the Effectiveness of Environmental Dredging
Projects, 72
Compilations and Reviews of Sediment Remediation Projects, 73
Criteria Used To Select Environmental-Dredging Projects for Evaluation, 75
Measures of Sediment-Remedy Effectiveness in the Superfund Program, 85
Statistical Analysis of Fish PCB Body Burdens, Grasse River, New York, 117
Correlations between Suspended Solids and Contaminant Concentrations, 120
Statistical Analysis of Mercury Concentrations in Surficial Sediment, Lavaca
Bay, Texas, 124
Description of the Substrate Topography of the Grasse River During
Dredging, 126
Statistical Analysis of Surficial Sediment PCB Concentrations, Grasse
River, New York, 127
Statistical Analysis of PCB Concentrations in Fish, Waukegan Harbor,
Illinois, 146
Estimated Monitoring Costs for Lower Fox River and Green Bay,
Wisconsin, ROD Remedy, 182
Common Physical, Chemical, and Biological Measurements Used To
Characterize Contaminated Sediments, 186
Monitoring of Conditions During Hot-spot Dredging at the New Bedford
Harbor Superfund Site for Effects on Human Exposure, 199
Verification Sampling at Harbor Island, Washington, 203
Delineating the Dredge Prism in the Fox River, Wisconsin, 204
Verification Sampling of Dredging Residuals at the Head of Hylebos Site,
Commencement Bay, Washington, 205
Dredging and Later Sedimentation at Manistique Harbor, Michigan, 208
Collecting Aquatic Samples for Monitoring Human Exposure, 209
Biodynamic Modeling to Predict Organism PCB Concentrations, 221
Six-Step Adaptive-management Process, 246
Examples of the Application of Adaptive-management Principles in
Sediment Remediation, 247
Importance of Quantifying Uncertainty in Risk Estimates, 254
FIGURES
2-1
Generic conceptual site model indicating contaminated sediment exposure
pathways between sediment and ecologic receptors, including fish, shellfish,
benthic invertebrates, birds, and mammals, 35
xviii
2-2
2-3
2-4
2-5
2-6
2-7
3-1
4-1
4-2
4-3
4-4
4-5
4-6
4-7
4-8
4-9
4-10
4-11
4-12
4-13
4-14
4-15
Contents
Historical changes in sediment core profiles of mercury concentrations collected
from Bellingham Bay, WA, during natural recovery (dredging or capping was
not performed), 36
Categories of dredging and sediment removal equipment, 52
Dredging-process train, 53
Cost breakdown of components of environmental dredging at the New Bedford
Harbor (New Bedford, MA) project in 2004, 55
Cost breakdown of components of environmental dredging at the head of the
Hylebos Waterway (Commencement Bay, WA), 55
Conceptual illustration of environmental dredging and processes, 58
Locations of environmental-dredging projects selected for evaluation, 82
Prevalence of cancer and pre-cancerous lesions in Black River bullheads before
and after dredging, 99
Map showing the Marathon Battery Superfund site on the Hudson River, NY,
adjacent to Cold Spring, NY, 100
Median sediment cadmium concentrations, East Foundry Cove, Marathon
Battery Site, NY, 102
Log probability plot of sediment cadmium concentrations, East Foundry Cove,
Marathon Battery, NY, 103
Ratio of cadmium body burdens in various animals and plants after dredging
to pre-dredging levels for the period 1996-2000, Marathon Battery Site, NY, 104
Sediment or contaminant mass resuspended or left as a residual, 108
Estimated generated residuals for 11 projects with data broken down on
whether debris, rock, or hardpan was prevalent at the site, 110
Average water-column PCBs and TSS during debris and sediment removal
activities (conducted simultaneously) in the Grasse River during 2005
demonstration dredging program, 113
PCB concentrations in water samples collected approximately 0.5 miles
downstream of the dredging operations in the Grasse River (NY), 114
Lipid-normalized PCB concentration in resident spottail shiners in Grasse River,
Massena, NY, 115
PCB concentrations in fish collected at the Duwamish Diagonal site pre- and
post-dredging, 119
Left, mean and 95% confidence intervals of surficial mercury concentrations
(mg/kg) (log base 10 scale) before dredging and after each of four dredging
passes. Right, concentration distribution and sample sizes (N) of surficial
mercury stratified by geographic region—Capa (CA), North Capa (NC), AA,
and the Trench Wall (TW)—and dredging pass (Pre, 1, 2, 3, 4; pre = predredging), 125
Distribution and sample sizes for Grasse River surficial sediment PCB
concentrations (mg/kg dry weight) stratified by geographic region, 128
Upstream to downstream transects of the Grasse River remediation area
showing: elevations of probing based cut line (dotted lines) based on the
pre-dredging (Spring 2005) estimated depth of contamination; post-dredging
(Fall 2005) extent of contamination (diamonds) determined by vibracore
sampling; and the post-dredging (Fall 2005) bottom elevation determined
by bathymetry, 142
Lipid-normalized PCB concentrations in carp in Waukegan Harbor, IL, 147
Contents
5-1
6-1
xix
Sediment profile imagery (SPI) equipment (two left photos) and sediment
profile photograph (right) from New Bedford Harbor Superfund site, 187
EPA’s tiered approach to the use of probabilistic risk assessment (PRA), 253
TABLES
2-1
3-1
4-1
4-2
5-1
Sediment Megasites (Sites at Which Remediation of the Sediment Component
Is Expected To Be at Least $50 million), 30
Summary of 26 Environmental-Dredging Projects Selected for Evaluation, 77
Summary of Pre-dredging and Post-dredging Verification Sampling Results
(2005) from Three Subunits of Operable Unit 1 in the Lower Fox River, 134
Spottail Shiner PCB Concentrations After Remediation of the GM Massena, NY,
Site in 1995, 150
Strengths and Limitations of Methods for Assessing Biologic Effects in Aquatic
Ecosystems, 190
Summary
BACKGROUND
Contaminated sediments in aquatic environments can pose risks to
human health and other organisms. Nationwide, the full extent of the
problem is poorly documented, but it is well known that many rivers,
harbors, lakes, and estuaries fed by existing or former industrial, agricultural, or mining areas contain contaminated sediments. It is also well
known that contaminants in the sediments can directly harm aquatic organisms or accumulate in their tissue, which can be consumed by humans. The potential adverse effects on human health and the environment are compelling reasons to reduce such exposures.
From a regulatory standpoint, contaminated sediments are challenging to manage. The Superfund program,1 administered by the U.S.
Environmental Protection Agency (EPA), is intended to protect human
health and the environment from hazardous substances at contaminated
sites. At most contaminated sediment Superfund sites, the remedial
The Superfund requirements are set forth in the Comprehensive Environmental Response, Compensation, and Liability Act (CERCLA) of 1980 as
amended (42 USC §§ 9601-9675 [2001]), and its implementing regulations are set
forth in the National Contingency Plan (NCP) (40 CFR 300).
1
1
2
Sediment Dredging at Superfund Megasites
process includes a site investigation, comparison of remedial alternatives, and selection and implementation of a remedy. The process is affected by numerous scientific and technical issues. Resource-intensive
surveys and analyses are required to document the distribution, depth,
and concentration of contaminants in sediments and contaminant concentrations in the aquatic biota. Even if substantial resources are focused
on a relatively small site, understanding the current and potential risks
of contaminated sediments can be difficult and uncertain, relying heavily
on surrogate measures and modeling of actual environmental effects.
The process of estimating and comparing the modeled results of potential remedial actions (including no action) has substantial uncertainties
that depend on a host of variables, including whether environmental
conditions have been adequately characterized and the accuracy of nearterm and long-term predictions of post-remediation contaminant behavior. The uncertainties are magnified by increasing duration of a remedial
action and increased extent and complexity of a contaminated site.
Contaminated sediments exist in a variety of environments and can
differ greatly in type and degree of contamination. Site conditions are
important in determining which remediation techniques (and combinations thereof) are appropriate. The techniques include removing the
sediments from the aquatic environment (for example, by dredging),
capping or covering contaminated sediments with clean material, and
relying on natural processes while monitoring the sediments to ensure
that contaminant exposures are decreasing, or at least not increasing.
Those approaches differ in complexity and cost; dredging is the most
complex and expensive, and monitoring without active remediation is
the least difficult and least expensive. Remedial approaches have tradeoffs with respect to the risks that are created during implementation and
that remain after remediation. Dredging may create exposures (for example, through the resuspension of buried contaminants) during implementation, but it has the potential to remove persistent contaminants
permanently from the aquatic environment. Monitoring without removal
does not itself create risks, but it leaves contaminants in the aquatic environment. Remedial operations also vary in efficacy within and among
the different approaches. The variability is driven by several factors, including site conditions and implementation of the remedial approach.
Decisions about whether to dredge at contaminated sediment sites
have proved to be among the most controversial at Superfund mega-
Summary
3
sites.2 The scientific and technical difficulties described above are augmented by the challenges of implementing a program that holds responsible parties liable for paying for the cleanup, with parties in some cases
unwilling to accept the liability. Cleanup planning must also be responsive to the public, which may have little tolerance for remedial actions
that leave contaminants in the local environment. Those controversies
often expand with the magnitude of the sites and the scope of remedial
activity, as has been seen at some of the nation’s largest sediment remediation sites (for example, the Fox River, WI, and the Hudson River, NY).
Regardless of cost or controversy, achieving the intended effect of
remedial actions in terms of anticipated improvements in the environment is of primary importance. That is true for regulators who require
cleanup of a site, parties responsible for funding the cleanup, and communities and user groups, such as anglers and boaters, who are directly
affected by the contamination.
THE COMMITTEE’S CHARGE
This study, which was conducted to evaluate the effectiveness of
dredging as a remedial option at contaminated sediment sites, originated
in the fiscal year 2006 appropriations bill for the Department of Interior,
environment, and related agencies. The accompanying conference report
(Report 109-188) states that “the managers believe that the appropriate
role for the NAS [National Academy of Sciences] is to act as an independent peer review body that will conduct an objective evaluation of
some of the ongoing dredging projects underway at Superfund
megasites. By undertaking such an evaluation, the Academy can serve as
an objective voice on this issue.”
The Committee on Sediment Dredging at Superfund Megasites was
convened by the National Research Council of the National Academies.
In brief, the committee’s charge requests an evaluation of the expected
effectiveness of dredging of contaminated sediments at Superfund
megasites. The committee was asked to consider such aspects of dredging as short-term and long-term changes in contaminant transport and
ecologic effects. Overall, the committee was charged to strive to develop
Megasites are those Superfund sites where remedial expenditures are expected to exceed $50 million.
2
4
recommendations that would facilitate scientifically based and timely
decision-making for megasites in the future but not to recommend particular remedial strategies for specific sites.
The committee, chosen by the National Research Council, consists
of experts in a variety of fields relevant to the statement of task. Over the
course of the study, the committee held three public sessions in which it
heard presentations on an array of dredging projects and received input
from members of the public and other interested parties; two closed, deliberative sessions were also held over the course of the year-long study.
EVALUATION OF EXPERIENCE AT DREDGING SITES
The committee examined experience at 26 dredging projects and
evaluated whether, after dredging, the cleanup goals had been met. That
involved evaluating whether a site achieved its quantitative cleanup levels (typically a specified concentration in sediment expected to be
achieved in the short term, that is, immediately after remedy implementation) and remedial-action objectives (often a narrative statement of
what the cleanup is expected to accomplish in the long term).3 Various
sites were examined, including full-scale dredging projects, pilot studies
at sites, and dredging projects within a large-scale remediation effort.
Conclusions
The committee concluded that dredging is one of the few options
available for the remediation of contaminated sediment and that it
should be considered, with other options, to manage the risks that the
contaminated sediments pose. However, the committee could not generally establish whether dredging alone is capable of long-term risk reduction. That is because monitoring at most sites does not include all the
In this context, short term refers to anything caused as an immediate consequence of the action being focused on; long-term extends beyond this time period
to the time required to achieve remedy success. For dredging projects, this would
include any period of natural recovery that is part of the remedy and necessary
to achieve the goals of the remedy.
3
Summary
5
BOX S-1 Overview of Conclusions and Recommendations
Sediment dredging can be effectively implemented to remove contaminants from aquatic systems, but technical limitations often constrain its ability to
achieve expected outcomes. The range of experiences and outcomes at dredging
sites, coupled with shortcomings in monitoring data, the lack of sufficient time to
observe long-term changes, and difficulties in separating the effects of dredging
from the effects of other processes limited the committee’s ability to establish
whether dredging alone has been effective in risk reduction. However, assessment of data from dredging projects does indicate that dredging has encountered
systematic difficulties in achieving specified cleanup levels (expected sedimentcontaminant concentrations after dredging) and that monitoring to evaluate
long-term success is generally lacking. The inability to meet desired cleanup levels is associated primarily with “residual” contamination that typically results
from dredging operations or from leaving contaminated sediment exposed after
dredging. Site assessments also indicate that contaminants can be released into
the water during dredging and can have short-term adverse effects on the
aquatic biota. Residual contamination and contaminant release are inevitable
during dredging and should be explicitly considered in estimating risk reduction. Some site conditions and dredging practices can limit the amount of residual contamination remaining after dredging and can limit contaminants released
into the water column. Those site conditions should be given major consideration
when evaluating the potential effectiveness of dredging.
Environmental monitoring is the only way to evaluate remedial success,
but monitoring at most Superfund sites has been inadequate to determine
whether dredging has been effective in achieving remedial objectives (that has
not been the case in several highly monitored pilot studies). Basic information
was not collected at some sites, and others had only recently completed dredging, so long-term trends could not be assessed. EPA should ensure that adequate
monitoring is conducted at all contaminated sediment megasites to evaluate remedial effectiveness. Some current monitoring techniques have proved useful in
determining short-term and long-term effects of remediation, but further development of monitoring strategies is needed. Pre-remediation monitoring is necessary to adequately characterize site conditions and to assemble a consistent longterm dataset that allows statistically valid comparisons with future postremediation monitoring data. Monitoring data should also be made available to
the public in an accessible electronic form so that evaluations of remedial effectiveness can be independently verified.
(Continued on next page)
6
BOX S-1 Continued
Regarding the future practices and management of contaminated sediments at megasites, the committee recommends that adaptive-management approaches should be implemented in the selection and implementation of remedies where there is a high degree of uncertainty about the effectiveness of the
remedial action. In selecting site remedies, dredging remains one of the few options available for the remediation of contaminated sediment, and it should be
considered along with other options, but EPA should compare the expected net
risk reduction associated with each remedial alternative, taking into account the
range of risks and uncertainty associated with each alternative. EPA should centralize sediment-related resources, responsibility, and authority at the national
level to ensure that necessary improvements are made in site tracking, in the
implementation of monitoring and adaptive management, and in research to
examine the relationship between the remedial actions, site conditions, and risk
reduction.
measures necessary to evaluate risk over time,4 dredging may have occurred in concert with other remedies or natural processes that affect
risk, insufficient time has passed to evaluate long-term risk reduction,
and a systematic compilation of site data necessary to track remedial effectiveness nationally is lacking.
However, the committee was able to draw several conclusions and
derive recommendations on the basis of monitoring data from a range of
dredging projects and by evaluating factors that affected their success.
The analysis indicates that dredging can be effective for removal of mass,
but that mass removal alone does not necessarily achieve risk-based
goals. Monitoring data demonstrate that dredging can have short-term
adverse effects, including increased contaminant concentrations in the
water, increased contaminant concentrations in the tissues of caged fish
adjacent to the dredging activity, and short-term increases in tissue contaminant concentrations in other resident biota. However, monitoring for
those effects was not conducted at many sites.
Monitoring for risk reduction is not straightforward; there is no analytic determination of “risk,” and estimating risk reduction requires multiple metrics
(for example, concentration, toxicity and bioavailability data) to be measured
consistently through time.
4
Summary
7
The most frequent post-dredging measurement used to assess effectiveness at the sites was contaminant concentrations in surface sediment. Surface concentrations (as opposed to concentrations in deeply
buried sediments) are the most relevant to risk. At some sites (for example, Bayou Bonfouca, LA; Waukegan Harbor, IL; and the Dupont Newport Site, Christina River, DE), sediments were not sampled for contaminants immediately following dredging. The committee’s analysis of predredging and post-dredging surface sediment concentrations indicates a
wide range of outcomes: some sites showed increases, some no change,
and some decreases in contaminant concentrations. Residual contamination after dredging can result from the incomplete removal of targeted
sediments or the deposition of sediment resuspended during dredging.
Residual contaminated sediments hamper the ability to achieve desired
cleanup levels and are exacerbated by site conditions like obstructions in
the dredging area and impenetrable or uneven formations underlying
the contaminated sediments. Overall, the committee found that dredging
alone achieved the desired contaminant-specific cleanup levels at only a
few of the 26 dredging projects, and that capping5 after dredging was
often necessary to achieve cleanup levels.
The committee was able to identify factors that led to the success or
failure of projects to meet desired short-term cleanup levels and, where
long-term data were available, remedial-action objectives. Some sites
exhibited conditions that are more conducive to dredging and less prone
to releasing contaminants and less likely to result in residual contaminated sediment after dredging. Favorable site conditions include little or
no debris (for example, rocks, boulders, cables, automobiles, and Ibeams), sediment characteristics that permit rapid (even visual) determination of clean vs contaminated sediment, conditions that allow overdredging into clean material beneath contaminated sediment (sites underlain by bedrock or hardpan are highly problematic), low-gradient
bottom and side slopes, lack of piers and other structures, rapid natural
attenuation processes after dredging, and absence of contaminants that
distribute to the water column rapidly after sediment disturbance.
Capping refers broadly to the placement of a layer of uncontaminated material over material with elevated concentrations to contain contaminated sediment.
5
8
The design and implementation of remediation can also influence
the extent of chemical release and residual contamination (as well as
counteract the influence of poor site conditions). Adequate site characterization is needed to identify adverse conditions and potential sources
of recontamination in the site or watershed. Pilot studies are particularly
useful for identifying adverse site conditions and logistical problems. As
discussed in the report, best-management practices during dredging can
help control residuals and resuspension. Backfilling and capping can be
used following dredging to manage residual contaminated sediments.
Contracting mechanisms can be used to encourage a focus on specified
cleanup levels and remedial-action objectives instead of on attaining
mass removal targets.
The combined experience indicates that dredging alone is unlikely
to be effective in meeting short-term and long-term goals if a site has one
or more unfavorable conditions. If conditions are unfavorable, contaminant resuspension and release and residual contamination will tend to
limit the ability to meet desired cleanup levels and will delay the
achievement of remedial-action objectives unless managed with a combination of remedies.
Recommendations
• Remedies should be designed to meet long-term risk-reduction
goals (as opposed to metrics not strictly related to risk, such as massremoval targets). The design should be tested by modeling and monitoring the achievement of long-term remedial action objectives.
• Environmental conditions that limit or favor the effectiveness of
dredging should be given major consideration in deciding whether to
dredge at a site.
• Resuspension, release, and residuals will occur if dredging is
performed. Decision-making should include forecasts to estimate the
effects of those processes, and the predictions should be explicitly considered in expectations of risk reduction. To reduce adverse effects, bestmanagement practices that limit resuspension and residual contamination should be used during dredging. The ability of combination remedies to lessen the adverse effect of residuals should be considered.
Summary
9
• Further research should be conducted to define mechanisms,
rates, and effects of residuals and contaminant resuspension associated
with dredging. It is known that contaminated sediment resuspension
and residuals create exposures, but the relationship of the magnitude of
those processes to environmental conditions, operational controls, and
management practices is not well quantified.
MONITORING FOR EFFECTIVENESS
Conclusions
Environmental monitoring is the only way to evaluate a remedy’s
success in reducing risk and ensure that the objectives of remediation
have been met. It is therefore an essential part of the remedy. It is impossible to evaluate effectiveness in the absence of sufficient baseline data
and appropriate reference sites. Monitoring needs to be conducted to
confirm not only that desired cleanup levels have been met, but that they
result in risk reduction.
In most cases reviewed by the committee, monitoring was designed
or implemented inadequately to determine whether dredging was effective in achieving the objectives of the remediation or long-term risk reduction. For example, at some sites, sparse or incomplete monitoring
data were collected. Pre-remediation monitoring approaches were not
always consistent with those used for long-term post-remediation monitoring. Pre-remediation trends in sediment or fish concentrations were
not of sufficient duration to enable judging the effect of the remedial action. The models and forecasts used to select a remedy were not updated
with post-remediation data to determine whether remediation had the
expected effects (or to examine why or why not). Monitoring was of insufficient quality or quantity to support rigorous statistical analyses.
However, some monitoring techniques have proved useful in determining short-term and long-term effects of remediation. Monitoring
during dredging, including measurements of mass flux (contaminant
transport over time) attained through upstream and downstream chemical monitoring, and biologic monitoring, including caged-fish studies
10
and passive sampling devices,6 are useful in indicating chemical releases
to the water from dredging. Chemical concentrations do not always correspond directly with potential uptake or toxicity because contaminant
bioavailability can differ among sampling locations. However, chemical
analyses are among the most highly standardized and easily attainable
indicators of risk and are useful in evaluating trends before and after
dredging. Laboratory toxicity testing has proven valuable in long-term
monitoring of risk to benthic organisms in sediment following dredging.
Fish-tissue monitoring for contaminants is useful for indicating
risks to the health of people and other piscivorous species, particularly if
long-term monitoring trend data exist. However, linking changes in fishtissue concentrations to remedial actions can be problematic, because
fish are mobile and can be exposed to offsite conditions. For describing
possible ecologic effects, benthic organisms (or organisms with home
ranges limited to the site) or passive sampling devices are probably better indicators, although not necessarily sufficient indicators of exposure
to higher trophic level species.
Recommendations
• EPA should ensure that monitoring is conducted at all contaminated sediment megasites to evaluate remedy effectiveness. Monitoring
data should be made available to the public in a form that makes it possible to verify evaluations of remedial efficacy independently.
• Monitoring plans should focus on elements required to judge
remedial effectiveness and to inform management decisions about a site.
Planning, evaluation, and adaptive management7 based on monitoring
findings should be closely linked to the conceptual site model so that the
hypotheses and assumptions that led to the selected remedy can be
tested and refined.
• Pre-remediation baseline monitoring methods and strategies
should be developed to allow statistically valid comparisons with postPassive sampling devices accumulate chemicals of interest during an extended deployment period (days to weeks) in the environment to provide an
integrated estimate of chemical exposure over that period.
7Adaptive management is generally used to learn from, test, assess, and modify or improve remedies with the goal of meeting long-term objectives.
6
Summary
11
remediation monitoring datasets. The ultimate goal is to assemble a consistent long-term dataset that can be used in evaluations. Monitoring
should be initiated during the design of the remedy to help to establish a
pre-remedial time trend, integrating earlier characterization data as
technically appropriate.
• Research in and development of rapid field monitoring techniques to inform dredging operations in nearly real time are needed to
indicate the effects of resuspension of contaminants and their release to
the water column. Biota monitoring approaches that use benthic invertebrates (or other organisms with appropriate home ranges) as indicators
of food-web transfer of contaminants should also be developed.
IMPROVING FUTURE DECISION-MAKING AT
SUPERFUND MEGASITES
Conclusions
The historical perspective and hindsight gained from the committee’s retrospective analysis of sediment sites provide an opportunity to
derive common lessons and to improve on the manner in which environmental dredging is planned and implemented. It is important that
this type of review be on-going and be part of a shared experience
among regulators, practitioners, and the public. The large spatial scale
and long remedial timeframes of contaminated sediment megasites make
it difficult to predict and quantify the human health and ecosystem riskreduction benefits achieved by isolated remediations in a large-scale watershed. In addition, the complexity and heterogeneity of large-scale
megasites suggest that a variety and combination of remedial approaches will often be appropriate. The committee concludes that three
critical kinds of changes are needed to improve decision-making and the
efficacy of dredging remedies at contaminated sediment megasites.
First, owing to the complexity, large spatial scale, and long time
frame involved, the management of contaminated megasites needs to
embrace a more flexible and adaptive approach to accommodate unanticipated factors, new knowledge, technology changes, and results of
field pilot tests. Comprehensively characterizing sediment deposits and
contaminant sources on the scale of megasites presents tremendous
12
technical challenges and uncertainties. Many large and complex contaminated sediment sites will take years or decades to remediate and will
encounter unforeseen conditions. A priori predictions of the outcome of
that remediation, made on the basis of incomplete information, will also
have high uncertainty. The typical Superfund approach, wherein EPA
conducts a remedial investigation and a feasibility study that establishes
a single path to remediation in the record of decision is not the best approach to remedy selection and implementation at these sites. At the
largest contaminated sediment sites, the remediation timeframes and
spatial scales are in many ways unprecedented. Remedial strategies will
often require unexpected adjustments, whether in response to new
knowledge about site conditions or advances in technology (such as improved dredge or cap design or in situ treatments). Because such uncertainty exists, regulators and others will need to adapt continuously to
evolving conditions and environmental responses that cannot be foreseen. Thus, the process for remedy selection at large, complex sediment
megasites needs to allow more adaptive site investigation, remedy selection, and remedy implementation.
Second, improved risk assessment that specifically considers the
full range and real-world limitations of remedial alternatives is needed
to allow valid comparisons of technologies and uncertainties. Each remedial alternative offers a unique set of risk-reduction benefits, possibly
with the creation of new contaminant exposure pathways and associated
risks. The effects of adverse environmental conditions, such as those
leading to chemical release and production of contaminated residuals,
need to be accounted for in a quantitative comparison of net risk reduction associated with different alternatives, and uncertainty should be
quantified to the extent warranted to optimize decision-making. Some
potential effects not related to chemical exposures (for example, quality
of life impacts and some risks to workers or the community from implementing a remedy) will remain difficult to quantify and to compare.
However, ignoring these risks in comparisons of remedial alternatives is
not the solution and may lead to undesirable consequences. Quantitative
assessment and comparison of some of these types of risks will probably
never be fully achievable, but they should be identified as risks associated with particular alternatives and considered at least qualitatively
when remedial alternatives are being compared and the risks and benefits associated with various options are presented to the public.
Summary
13
Third, EPA needs to centralize and coordinate assessment and
management of contaminated sediment megasites to ensure greater consistency in evaluations, greater technical competence, more active leadership at the sites, and an emphasis on what works and why. Several
specific responsibilities are discussed in the corresponding recommendation below.
Recommendations
• An adaptive-management approach is essential to the selection
and implementation of remedies at contaminated sediment megasites
where there is a high degree of uncertainty about the effectiveness of
dredging.
• Adaptive approaches based on the use of monitoring data from
pilot studies and remedial operations should be used to learn from, test,
assess, and modify or improve remedies with the goal of meeting longterm objectives.
• EPA should compare the estimated net risk reductions associated with different remedial alternatives, taking into account the realworld limitations of each approach (such as residuals and resuspension)
in selecting site remedies.
• EPA should centralize resources, responsibility, and authority at
the national level to ensure that necessary improvements are made so
that contaminated sediment megasites are remediated as effectively as
possible. Responsibilities would include
— Gathering data to define the scope of the contaminated sediment
problem nationally and track likely future contaminated sediment
megasites. Defining the scope of the contaminated sediment problem
is important for two reasons: this will help place the magnitude of
the problem in proper perspective by establishing how much of the
problem has been addressed and how much remains, and documenting remaining work and associated costs should help EPA and
Congress identify the most pressing program and research needs.
— Reviewing site studies, remedies, and monitoring approaches at
contaminated sediment megasites to assess whether best practices are being
implemented, including whether regions are complying with national sedi-
14
ment and other program guidance. Because every EPA region has onthe-ground management and remediation experience with dredging
at some megasites, regular review and shared experience can inform
decision-making and raise the overall level of technical expertise.
The goal is to generate a greater understanding of sound remediation principles and best practices and their consistent application
among sites.
— Ensuring that adaptive-management principles and approaches are
applied at contaminated sediment megasites. As described above, a
phased, adaptive approach will be required in remedy selection and
implementation at large, complex megasites. EPA should ensure that
monitoring data are used to support and update forecasts of the effects of remedial measures and to adapt a remedy if remedial goals
are not achieved in the expected time frame.
— Ensuring adequate pre-remediation and post-remediation monitoring and evaluating sediment cleanups in nearly real time to determine
whether remedies are having the intended effects. Without adequate predredging and post-dredging monitoring, it is impossible to evaluate
the degree to which cleanup has achieved remedial objectives. EPA
should invest in better and more consistent measurement tools to
monitor conditions in the field reliably and efficiently. EPA should
ensure that these techniques are used before and after remediation
so that the effectiveness of the projects can be assessed. To facilitate
information transfer, a centralized, easily accessible, and up-to-date
repository of relevant data and lessons learned regarding sediment
remedies should be created.
— Developing and implementing a research strategy, including new
technologies, for contaminated sediment sites. EPA and its federal partners should develop a research and evaluation strategy to understand the risk reduction attained by various technologies under various site conditions and the associated uncertainty. EPA’s efforts
should focus on moving forward with remedies at sites while testing
and learning with each new pilot test or remedy to determine what
works, what does not work, and why. Research to improve and develop new remediation technologies, site-characterization techniques, and monitoring tools is essential to advance sediment remediation and should be supported.
Summary
15
Many of the sites are vast and expensive, and it is worthwhile to
invest time and resources now to ensure more cost-effective remedies in
the future. That focus is warranted if the country is to make the best use
of the billions of dollars yet to be spent on remediation.
1
Introduction
THE CHALLENGE OF CONTAMINATED SEDIMENTS
Contaminated sediments in aquatic environments can pose health
risks to many types of organisms, including humans. Exposure to the
contaminants occurs by several routes, including direct contact and consumption of organisms that have accumulated contaminants from the
sediments. The potential adverse effects on human health and the environment are compelling reasons to seek to reduce exposure.
Contaminated sediments can occur in small, localized areas or in
vast areas, covering miles of river or harbor bottoms and associated
floodplains. They occur in wetlands, coastal tidal flats and embayments,
ocean basins, lakes, rivers, and streams. In some cases, contamination is
relatively contained; in other cases, contaminated sediment exists
throughout a watershed and may have multiple sources of contamination, including stormwater and sewer outfalls, industrial discharges, agricultural runoff, and atmospheric deposition.
The chemicals of concern in contaminated sediment sites vary;
polychlorinated biphenyls (PCBs) are the most common, followed by
metals, and polycyclic aromatic hydrocarbons (EPA 2005). The widely
varied physical and chemical properties of contaminants markedly affect
their distribution in the environment and their behavior (including
transport, bioavailability, and toxicity) during and after remediation. The
16
Introduction
17
degree of contamination can be severe in some areas with nearly unadulterated original products, such as PCB-containing oils, pesticides, or
coal-tar residues. In other areas, contaminants occur at low concentrations in sediments among functioning ecosystems of fish, plants, and
benthic invertebrates. The thickness of the contaminated sediment is
highly variable and often poorly characterized but can range from a few
inches to many feet thick with marked differences over small spatial
scales. In addition, the nature of the sediments and particularly of the
underlying substrate can vary widely on the basis of local geology, hydrology, and human activities that have altered the watershed characteristics.
Because of the highly variable nature of sediments, the environments in which they occur, and the type and degree of contamination,
there are many approaches to their remediation. The techniques, which
can be employed in combination, include removing the sediments from
the aquatic environment (for example, by dredging), capping or covering
contaminated sediments with clean material, and relying on natural
processes while monitoring the sediments to ensure that contaminant
exposures are decreasing, or at least not increasing (known as monitored
natural recovery [MNR]). In-situ treatments that, for example, reduce the
bioavailability of contaminants can also be used. The techniques, which
are examined in greater detail in Chapter 2, differ in complexity, cost,
efficacy, and time frame. That variability is driven by several factors, including site conditions (for example, variations in water flow and depth),
underlying substrate characteristics, and implementation of the remedial
approach. Regardless, achieving expected reduction in risk is of primary
importance to regulators who require cleanup of a contaminated sediment site, parties responsible for funding the cleanup, and communities
and user groups that are directly affected by the contamination and the
remediation process.
Managing the risks associated with contaminated sediments has
been an issue at the federal level since at least the middle 1970s (Johanson and Johnson 1976, as cited in EPA 1987), although it received substantially greater attention in the 1980s when the U.S. Environmental
Protection Agency (EPA) sought to document the nature and extent of
sediment contamination (Bolton et al. 1985; EPA 1987). The 1989 National Research Council report Contaminated Marine Sediments: Assessment
and Remediation (NRC 1989) examined the extent of and corresponding
18
risk posed by marine sediment contamination and examined remedial
technologies. EPA’s Assessment and Remediation of Contaminated
Sediments (ARCS) program—an early effort to understand the extent of,
associated risks of, and techniques for remediating contaminated sediments—published several useful reports and guidance documents dealing with the assessment of contaminated sediments and various treatment technologies (EPA 1994). Since then, additional National Research
Council reports on managing contaminated sediments have been released (NRC 1997, 2001), and EPA has published its sediment quality
surveys (EPA 1997; 2004), produced a contaminated sediment management strategy (EPA 1998), and issued comprehensive contaminated
sediment guidance (EPA 2005). Yet, even after decades of analysis and
review, managing and remediating contaminated sediments remains a
major scientific and management challenge. Areas with contaminated
sediments continue to be identified, and remediation efforts are increasingly large, expensive, and resource-intensive.
This report is one piece of the continuing dialogue and seeks to assess the effectiveness of environmental dredging for reducing risks associated with contaminated sediments, particularly at large, complex sites.
Environmental dredging is of special interest because it can be expensive
and technically challenging to implement. Dredging itself may create
exposures (for example, through the resuspension of buried contaminants), but it removes persistent contaminants (and their associated potential for transport and risk) from the aquatic environment permanently. Whether to dredge contaminated sediments has proved to be one
of the most controversial aspects of decision-making at sediment remediation sites.
THE CHARGE TO THE COMMITTEE ON SEDIMENT
DREDGING AT SUPERFUND MEGASITES
This study was requested in the fiscal year 2006 appropriations bill
for the Department of Interior, environment, and related agencies. The
accompanying conference report (Report 109-188) states that “the managers believe that the appropriate role for the NAS [National Academy
of Sciences] is to act as an independent peer review body that will conduct an objective evaluation of some of the ongoing dredging projects
Introduction
19
underway at Superfund megasites. By undertaking such an evaluation,
the NAS can serve as an objective voice on this issue.”1
In response, the National Research Council of the National Academies convened the Committee on Sediment Dredging at Superfund
Megasites to consider the specific tasks provided in the statement of task
(see Appendix A). In brief, the committee’s charge requests an evaluation of the expected effectiveness of dredging of contaminated sediments
at Superfund megasites and of whether risk-reduction benefits are expected to be achieved in the expected period. The committee was asked
to consider such aspects of dredging as the short- and long-term changes
in contaminant transport and ecologic effects. The statement of task also
directs the committee to evaluate monitoring strategies and whether
those strategies are sufficient to inform assessments of effectiveness.
Overall, the committee was charged to strive to develop recommendations that would facilitate scientifically based and timely decisionmaking for megasites in the future but not to recommend particular remedial strategies at specific sites.
One subject of great interest and concern at contaminated sediment
Superfund sites is the risk-based comparison of remedial alternatives
and selection of a remedy (Bridges et al. 2006; Wenning et al. 2006; Zeller
and Cushing 2006). The committee briefly discusses this topic (Chapter
2) and addresses it in the context of improving future decision making at
Superfund megasites (Chapter 6). However, the report does not develop
specific procedures and recommendations for performing comparative
risk analyses in selection of a sediment remedy. While that topic and
type of analysis is quite important, it was not requested of the committee
and it has not been undertaken.
NATIONAL RESEARCH COUNCIL AND
THE COMMITTEE PROCESS
The National Research Council is a nonfederal, nonprofit institution that provides objective science, technology, and health-policy advice
generally by producing consensus reports written by committees. It exists to provide independent advice; it has no government affiliation and
Megasites are those Superfund sites where remedial expenditures are expected to exceed $50 million.
1
20
no regulatory role. There is no direct oversight of a committee by the
study sponsor or any other outside parties. Thus, EPA and other interested parties have no more input or access to committee deliberations
than does the general public. That arrangement gives the committee
complete independence in conducting its study. The committee members
have a wide variety of backgrounds and expertise. Members are selected
by the Research Council primarily for their academic credentials and
their knowledge, training, and experience relevant to the statement of
task (see Appendix B for committee-member biosketches). Members
conduct their work solely as a public service, volunteering to the Research Council and the nation, cognizant of the importance of providing
timely and objective scientific advice.
In conducting its review and evaluation, the committee relied on
the Superfund-site decision documents and supporting materials, other
scientific studies, technical presentations made to the committee, other
information submitted by individuals and interest groups, and the committee’s observations and personal expertise. All information received by
Research Council staff that was made available to committee members is
available to the public through the Research Council’s public-access records office.
The committee held five meetings. Three included open, information-gathering sessions in which the committee heard from invited
speakers and from interested members of the public. The first meeting
(in March 2006) was in Washington, DC; the second was held in Irvine,
CA (June 2006); and the third was in Woods Hole and New Bedford, MA
(July 2006), where the committee toured an active dredge site and sediment-handling facility. All of the public meetings included an open session where anyone was able to provide comment to the assembled committee. In addition, the committee was available to receive written
materials throughout the study. The fourth and fifth meetings, held in
September and October 2006 in Washington, DC, were closed, deliberative sessions attended only by committee members and staff.
REPORT ORGANIZATION
Chapter 2 provides background on sediment management at
Superfund megasites; it includes discussion of the concept of reducing
Introduction
21
risk through environmental remediation and details on remedial techniques, particularly dredging. Chapter 3 describes the committee’s approach to considering effectiveness at various sites and developing conclusions from the analyses. Chapter 4 evaluates remedy performance and
risk reduction on the basis of sites’ pre-dredging and post-dredging
monitoring data and evaluates factors that affected performance. Chapter 5 looks at current practices for monitoring effectiveness at sediment
remediation sites and considers the types of assessments and protocols
that are needed to improve monitoring. Chapter 6 looks to the future: it
considers the implications of the committee’s assessment of sediment
management and identifies opportunities to advance the understanding
of dredging and its effectiveness in improving the environment and public health.
Overall, the committee recognizes that the state of the science of
environmental dredging is continually changing. New information is
being gathered, research detailing the effects and effectiveness of dredging is being conducted, and technologies and performance continue to
evolve. That process will continue for the foreseeable future. The committee does not consider its review to be the last word, but it hopes that
its findings and recommendations will assist government agencies and
other stakeholders in improving the approaches to contaminated sediments at large, complex megasites.
REFERENCES
Bolton, H.S., R.J. Breteler, B.W. Vigon, J.A. Scanlon, and S.L. Clark. 1985. National Perspective on Sediment Quality. EPA Contract No. 68-01-6986. U.S.
Environmental Protection Agency, Criteria and Standards Division, Office
of Water Regulations and Standards, Washington, DC.
Bridges, T.S., S.E. Apitz, L. Evison, K. Keckler, M. Logan, S. Nadeau, and R.J.
Wenning. 2006. Risk-based decision making to manage contaminated
sediments. Integr. Environ. Assess. Manage. 2(1):51-58.
EPA (U.S. Environmental Protection Agency). 1987. An Overview of Sediment
Quality in the United States. EPA-905/9-88-002. Office of Water Regulations and Standards, U.S. Environmental Protection Agency, Washington,
DC, and Region 5, Chicago, IL. June 1987.
EPA (U.S. Environmental Protection Agency). 1994. Assessment and Remediation of Contaminated Sediments (ARCS) Program: Final Summary Report.
22
EPA 905-S-94-001. Great Lakes National Program Office, U.S. Environmental Protection Agency, Chicago, IL. August 1994.
EPA (U.S. Environmental Protection Agency). 1997. The Incidence and Severity
of Sediment Contamination in Surface Waters of the United States: Report
to Congress, Vols. 1-3. EPA 823-R-97-006, EPA 823-R-97-007, EPA 823-R-97008. Office of Science and Technology, U.S. Environmental Protection
Agency, Washington, DC. September 1997 [online]. Available: http://
www.epa.gov/waterscience/cs/congress.html [accessed Dec. 26, 2006].
EPA (U.S. Environmental Protection Agency). 1998. EPA’s Contaminated Sediment Management Strategy. EPA-823-R-98-001. Office of Water, U.S. Environmental Protection Agency. April 1998 [online]. Available: http://www.
epa.gov/waterscience/cs/strategy.pdf [accessed Dec. 26, 2006].
of Sediment Contamination in Surface Waters of the United States:
National Sediment Quality Survey, 2nd Ed. EPA-823-R-04-007. Office of
Science and Technology, U.S. Environmental Protection Agency,
Washington, DC. November 2004 [online]. Available: http://www.epa.gov/
waterscience/cs/report/2004/nsqs2ed-complete.pdf [accessed Dec. 26, 2006].
EPA (U.S. Environmental Protection Agency). 2005. Contaminated Sediment
Remediation Guidance for Hazardous Waste Sites. EPA-540-R-05-012.
OSWER 9355.0-85. Office of Solid Waste and Emergency Response, U.S.
Environmental Protection Agency. December 2005 [online]. Available:
http://www.epa.gov/superfund/resources/sediment/pdfs/guidance.pdf [accessed Dec. 26, 2006].
Johanson, E.E., and J.C. Johnson. 1976. Identifying and Prioritizing Locations for
the Removal of In-Place Pollutants. Contract No. 68-01-2920. Office of Water Planning and Standards, U.S. Environmental Protection Agency, Washington, DC. May 1976.
NRC (National Research Council). 1989. Contaminated Marine Sediments: Assessment and Remediation. Washington, DC: National Academy Press.
NRC (National Research Council). 1997. Contaminated Sediments in Ports and
Waterways. Washington, DC: National Academy Press.
NRC (National Research Council). 2001. A Risk-Management Strategy for PCBContaminated Sediments. Washington, DC: National Academy Press.
Wenning, R.J., M. Sorensen, and V.S. Magar. 2006. Importance of implementation
and residual risk analyses in sediment remediation. Integr. Environ. Assess. Manage. 2(1):59-65.
Zeller, C., B. Cushing. 2006. Panel discussion: Remedy effectiveness: What works,
what doesn’t? Integr. Environ. Assess. Manage. 2(1):75-79.
2
Sediment Management at
Superfund Megasites
A variety of subjects including environmental engineering, toxicology, environmental monitoring, human and environmental risk assessment, and risk management are relevant to evaluating remediation at
contaminated sediment Superfund sites. In this chapter, a number of issues are briefly introduced to provide background for later discussions.
Topics include the Superfund process and information available on contaminated sediment Superfund sites; evaluating and managing risks
posed by contaminated sediments; and techniques for managing and
remediating contaminated sediment with a focus on dredging technologies and their performance capabilities and limitations. The chapter is
intended to provide a cursory overview of the topics while emphasizing
other sources containing more detailed discussions.
OVERVIEW OF SUPERFUND AND SEDIMENT MEGASITES
Superfund and Environmental Remediation
In 1980, Congress enacted the Comprehensive Environmental Response, Compensation, and Liability Act (CERCLA, 42 U.S.C. 9601-9675),
23
24
which authorized the establishment of the Superfund program. The goal
of the program is to reduce current and future risks to human health and
the environment at sites contaminated with hazardous substances.
CERCLA established a wide-ranging liability system that makes those
responsible for the contamination at sites liable for cleanup costs (see
Probst et al. 1995 for greater detail). It also created the “Superfund,” a
trust fund stocked primarily by a dedicated tax on oil and chemical
companies, to fund cleanup activities where there was no financially viable responsible party. Since the taxing authority expired in 1995, the
trust fund is largely depleted, and Congress now funds the program
from general revenues through annual appropriations (Fletcher et al.
2006).1 The U.S. Environmental Protection Agency (EPA) implements the
program through the National Oil and Hazardous Substances Pollution
Contingency Plan (40 CFR § 300), commonly referred to as the NCP or
the national contingency plan.
Most of the Superfund program’s efforts are aimed at cleaning up
sites on the National Priorities List (NPL). Typically, a site is proposed
for inclusion on the NPL after being evaluated with a hazard-ranking
system, which assesses the potential for hazardous-substance releases at
a site to harm human health or the environment (40 CFR § 300 Appendix
A). The Superfund process progresses from an initial site assessment
through cleanup and eventually deletion of the site from the NPL. Site
activities can be paid for by EPA (known as “fund-led” cleanups),2 by
parties connected to the site (referred to as responsible parties), or by
some combination of the two.
Selection of a remedy begins with a remedial investigation and feasibility study (RI/FS). The RI is intended to determine the nature and extent of contamination and estimate the associated risk to people and the
environment. The FS analyzes and compares remedial alternatives according to the nine NCP criteria (Box 2-1). The criteria require that the
remedy, above all, be protective of human health and the environment
and comply with all applicable or relevant and appropriate requirements
It is worth noting that, in the last few years, EPA has been in the position of
not having enough funds to fund all the new remedies that are ready to be
started at NPL sites (EPA 2004a).
2For fund-lead cleanups, states are required to pay 10% of the costs.
1
Sediment Management at Superfund Megasites
25
BOX 2-1 Evaluation Criteria for Superfund Remedial Alternatives
Before a remedial strategy is selected for a Superfund site, the options are
evaluated on the basis of nine criteria (see below). The first two, overall protection of human health and the environment and compliance with applicable or
relevant and appropriate requirements (ARARs), are termed threshold criteria,
and a potential remedy must meet them to be selected as a final remedy.3 The
next five criteria are termed balancing criteria and are used in weighing the advantages and disadvantages of potential remedies. The final two criteria are
modifying criteria, and the agency is supposed to take them into consideration as
part of the selection process.
Threshold Criteria
• Overall protection of human health and the environment. This criterion is
used to evaluate how the alternative as a whole achieves and maintains protection of human health and the environment.
• Compliance with applicable or relevant and appropriate requirements
(ARARs). This criterion is used to evaluate whether the alternative complies with
chemical-specific, action-specific, and location-specific ARARs or a waiver is
justified.
Balancing Criteria
• Long-term effectiveness and permanence. This criterion includes an evaluation of the magnitude of human health and ecologic risk posed by untreated contaminated materials or treatment residuals remaining after remedial action has
been concluded (known as residual risk) and of the adequacy and reliability of
controls to manage such risk. It also includes an assessment of the potential need
to replace technical components of the alternative.
• Reduction of toxicity, mobility, and volume through treatment. This criterion
refers to the evaluation of whether treatment processes can be used, the amount
of hazardous material treated (including the principal threat that can be addressed), the degree of expected reductions, the degree to which the treatment is
irreversible, and the type and quantity of treatment residuals.
• Short-term effectiveness. This criterion includes an evaluation of the effects of the alternative during the construction and implementation phase until
Except that specific ARARs can be waived.
3
26
BOX 2-1 Continued
remedial objectives are met. It includes an evaluation of protection of the community and workers during the remedial action, the environmental effects of
implementing the remedial action, and the expected length of time until remedial objectives are achieved.
• Implementability. This criterion is used to evaluate the technical feasibility of the alternative—including construction and operation, reliability, and
monitoring—and the ease of undertaking an additional remedial action if the
remedy fails. It also considers the administrative feasibility of activities needed
to coordinate with other offices and agencies—such as for obtaining permits for
off-site actions, rights of way, and institutional controls—and the availability of
services and materials necessary for the alternative, such as treatment, storage,
and disposal facilities.
• Cost. This criterion includes an evaluation of direct and indirect capital
costs, including costs of treatment and disposal; annual costs of operation, maintenance, and monitoring of the alternative, and the total present worth of these
costs.
Modifying Criteria
• State (or support agency) acceptance. This criterion is used to evaluate the
technical and administrative concerns of the state (or the support agency, in the
case of state-lead sites) regarding the alternatives, including an assessment of the
state’s or support agency’s position and key concerns regarding the alternative,
and comments on ARARs or the proposed use of waivers. Tribal acceptance is
also evaluated under this criterion.
• Community acceptance. This criterion includes an evaluation of the concerns of the public regarding the alternatives. It determines which component of
the alternatives interested persons in the community support, have reservations
about, or oppose.
Source: Adapted from EPA 2005a.
(ARARs).4 Remedies are also compared on whether they are technically
feasible and cost-effective, provide long-term (permanent) effectiveness,
and minimize deleterious effects and health risks during implementaARARs pertain to federal, state, or tribal environmental laws relevant to a
site.
4
27
tion. There is a preference for remedies that can reduce the toxicity, mobility, and volume of contaminants. Finally, there is a preference for
remedies that have state and community support.
EPA uses the FS to identify each alternative’s strengths and weaknesses and the trade-offs that must be balanced for the site in question
(EPA 1988). The agency then selects a remedy and describes it in a record
of decision (ROD). Additional studies may be conducted to support the
design of the remedy. Once constructed and implemented, the remedy is
maintained and monitored to ensure that it achieves its long-term goals.
EPA may delete a site from the NPL when a remedy has been implemented, the cleanup goals have been achieved, and the site is deemed
protective of human health and the environment (EPA 2000).
If, after implementation of a remedy, contamination exists that
could limit potential uses of the site, the site is subject to 5-year reviews
even if it has been deleted from the NPL (EPA 2001). The reviews are
intended to evaluate the performance of the remedy in protecting human
health and the environment and are to be based on site-specific data and
observations. However, monitoring is not limited to sites where 5-year
reviews are required. EPA guidance states that “most sites where contaminated sediment has been removed also should be monitored for
some period to ensure that cleanup levels and RAOs [remedial action
objectives] are met and will continue to be met” (EPA 2005a, p. 2-17).
Post-remediation monitoring (required in conjunction with 5-year reviews or otherwise) is the basis for evaluating remedy effectiveness and
adapting remedial strategies and risk management to achieve remedial
action objectives (for further discussion, see Chapter 5).
Sediment Contamination at Superfund Sites
Contaminated sediment is a widespread problem in the United
States (EPA 1994, 1997, 1998, 2004a, 2005a). Its wide distribution results
from the propensity of many contaminants discharged to surface waters
to accumulate in sediment or in suspended solids that later settle. Contaminants can persist in sediment over long periods if they do not degrade (for example, metals) or if they degrade very slowly (for example,
polychlorinated biphenyls [PCBs] or polycyclic aromatic hydrocarbons
28
[PAHs]). Historically contaminated sediment can become buried or, if it
is resuspended, can settle out eventually and lie on the sediment surface.
At the national level, the geographic extent of areas with contaminated sediment is not fully defined. In the 2004 Contaminated Sediment
Report to Congress (EPA 2004b), EPA reported on sediment sampling at
19,398 sampling stations nationwide, located in about 9% of the waterbody segments in the United States. Of that nonrandom sample of sediment sampling stations, EPA classified 43% as having probable adverse
effects, 30% having possible adverse effects, and 27% as having no indications of adverse effects. The 2005 Contaminated Sediment Remediation
Guidance for Hazardous Waste Sites (EPA 2005a) cites EPA fish advisories
covering all five Great Lakes, 35% of the nation’s other lakes, and 24% of
total river miles as due partly to sediment contamination (EPA 2005b).
EPA does not maintain a current list of NPL sites with contaminated sediments, nor does it compile a list of contaminated sediment areas that are potential Superfund sites. It also does not maintain a list of
contaminated sediment sites that are being (or have been) remediated
under another authority. EPA did report that “as of September 2005,
Superfund has selected a remedy at over 150 sediment sites” (EPA
2006a). In addition, the EPA Office of Superfund Remediation and Technology Innovation is tracking progress at 66 sites, termed tier 1 sites,
where the sediment-cleanup remedy involves more than 10,000 cubic
yards (cy) of sediment to be dredged or excavated or more than 5 acres
to be capped or monitored for natural recovery (EPA 2006b).5 Of the
aforementioned 150 NPL sites where remedies have been selected, EPA
considers 11 to be sediment megasites, defined as sites where the sediment portion of the remedy is expected to cost $50 million or more.6 Of
these 11 sites, 10 were proposed for inclusion on the NPL in the very
early years of the Superfund program (in 1982-1985), and one (Onondaga
Lake) was proposed for inclusion in 1993. Thus, the overwhelming ma-
The exact number of tier 1 sites is not clear. EPA’s website (EPA 2006b) lists
66 sites while 60 sites are listed in output from EPA’s internal database of tier 1
sites (EPA, unpublished data, “Remedial Action Objectives for Tier 1 Sites,” Sept.
5, 2006). Seven sites listed in EPA’s internal database are not on the website; 13
sites listed on the website are not in the September 5 submittal.
6Typically, megasites are defined as sites where the total cost of the remedy for
the entire site (not just the sediment portion) is expected to be at least $50 million.
5
29
jority of the megasites have been on the NPL for over 20 years. Only one
of the 11, Marathon Battery, has been formally deleted from the NPL. In
addition to the 11 megasites on the NPL, EPA lists two megasites that
have been proposed for the NPL but are not final (GE Housatonic River,
MA and Fox River, WI) and one that has not been proposed (Manistique
River/Harbor area, MI). The 14 sites are listed in Table 2-1. The status of
remediation at the sites varies. At some, such as Bayou Bonfouca and
Marathon Battery, remediation has been completed; at others, such as
Commencement Bay and Sheboygan Harbor, remedial activities are going on; and at still others, such as Hudson River and Onondaga Lake,
remedial activities have not begun. Megasites are described only in
terms of remediation cost (at least $50 million), so the size and volume of
contaminated materials at the sites can vary greatly (see Box 2-2).
One might ask, Why all this attention to contaminated sediment
megasites if there are only 14 nationwide? There are two reasons. First, at
13 of the megasites mentioned above (no cost information was provided
on the Triana/Tennessee River site), total remedial costs are estimated to
be about $3 billion, a huge amount of money even by Superfund standards.7 Second, the 14 sites probably constitute only a subset of the contaminated sediment sites that will entail expensive remedies and will be
cleaned up under the Superfund program. For example, the EPA list of
contaminated sediment megasites does not include some well-known
sites, such as the Bunker Hill Mining and Metallurgical Complex, ID,
and Love Canal, NY. Both those tier 1 sites are megasites by the conventional definition (total remediation cost of at least $50 million), but the
sediment portion alone is not expected to be $50 million.
When comparing EPA’s list of tier 1 sites (EPA 2006b) with a
somewhat dated list of megasites8 (that does not include federal facilities), one can find 11 “conventional” megasites on the tier 1 list.
• Alcoa–Point Comfort/Lavaca Bay, TX
• Allied Paper Inc./Portage Creek/Kalamazoo River, MI
• Bunker Hill Mining and Metallurgical Complex, ID
Based on data provided by EPA, “50M cost Query_091306.xls” (EPA, unpublished data, Sept. 18, 2006).
8Based on the report to Congress, Superfund’s Future: What Will It Cost? (Probst
et al. 2001), which lists megasites as of FY 2000.
7
30
TABLE 2-1 Sediment Megasites (Sites at Which Remediation of the Sediment
Component Is Expected To Be at Least $50 million)
Site Name, State
NPL Sites
New Bedford Harbor, MA
Hudson River PCBs, NY
Marathon Battery Corp., NY
Onondaga Lake, NY
Triana/Tennessee River, AL
Sheboygan Harbor and River, WI
Velsicol Chemical, MI
Bayou Bonfouca, LA
Milltown Reservoir Sediments, MT
Silver Bow Creek/Butte Area, MT
Commencement Bay, WA
Non-NPL Sites
GE Housatonic River, MA
Fox River, WI
Manistique River/Harbor area, MI
Source: EPA, unpublished data, “$50M Cost query_091306.xls,” Sept. 18, 2006.
•
•
•
•
•
•
•
•
Eagle Mine, CO
EI duPont–Newport landfill, DE
GM–Central Foundry Division (Massena), NY
Lipari landfill, NJ
Love Canal, NY
McCormick and Baxter Creosoting Co., CA
Nyanza chemical waste dump, MA
Wyckoff Co.–Eagle Harbor, WA
Furthermore, as described below, large and expensive sediment
remediations are conducted under authorities other than Superfund.
A crucial question is how many additional major contaminated
sediment sites are likely to be listed on the NPL. EPA does not designate
“likely future megasites” in its tier 1 list of sites or NPL sites for which
RODs have not been issued. According to EPA, the most likely future
sediment megasites are the “tier 2” contaminated sediment sites (S. Ells,
31
BOX 2-2 How Large Is a Megasite?
Contaminated sediment megasites are among the most challenging and
expensive sites on the NPL. Megasites are conventionally defined as those with
remedial activities costing at least $50 million, but there are large differences in
the magnitudes and scales of these sites. A few megasites, such as Bayou Bonfouca and Marathon Battery, are relatively small, with dredging activities covering tens of acres and operations occurring over a few years. Other dredging projects—such as those in the Fox River, New Bedford Harbor, and Commencement
Bay—are components of broader activities at large-scale megasites where remedial activities are going on and will take years or decades to complete. The $50million distinction for a megasite is not readily translatable into volume of materials removed. For example, sediment remediation (including design, mobilization, marine demolition, dredging, water management, transportation and disposal, construction oversight and EPA oversight, without the upland-based
removal costs) at the Head of Hylebos Waterway in Commencement Bay, WA,
removed 404,000 cy at a cost of $58.8 million (about $145/cy) (P. Fuglevand, personal commun., Dalton, Olmsted & Fuglevand, Inc., May 11, 2007). In Manistique Harbor, MI, dredging operations removed 187,000 cy at a cost of $48.2 million (about $260/cy) (Weston 2002). Dredging operations in Bayou Bonfouca, LA,
removed 170,000 cy at a cost of $90 million (about $530/cy) (EPA, unpublished
information, “$50M Cost query_091306.xls,” Sept. 18, 2006).
EPA, personal commun., Oct. 12, 2006). Tier 2 sites are designated for
review by the Contaminated Sediments Technical Advisory Group because they are large, complex, or controversial contaminated sediment
Superfund sites.9 There are 12 tier 2 sites. Three are on the earlier two
lists provided, but nine are not. Of the nine, four are NPL sites (Ashland/Northern States Power, WI, Portland Harbor, OR, Lower Duwamish Waterway, WA, and the Pearl Harbor Naval Complex, HI), and
five are not (Palos Verdes, CA, Kanawah River/Nitro, WV, Centredale
Manor Restoration Sites, RI, Anniston PCB site, AL, and Upper Columbia River, WA). EPA also indicates that the Passaic River, NJ, Berry’s
Creek at Ventron/Velsicol, NJ, and Tar Creek, OK, are likely future con-
Although, it should be noted that EPA indicates that “No quantifiable criteria
were used to develop this list.” The list of sites is available at http://www.epa.
gov/superfund/resources/sediment/cstag_sites.htm.
9
32
taminated sediment megasites, although they have not been designated
as tier 2 sites (S. Ells, EPA, personal commun., Sept. 18, 2006).
Because predicting future NPL listings is more an art than a science, in some ways, it is not surprising that there is no official list of
likely future contaminated sediment megasites. That said, the committee
was surprised that there is so little effort devoted to tracking and understanding likely future sediment megasites at the national level. Apparently, fewer than two full-time employees are assigned to contaminated
sediment issues at Superfund headquarters. It appears that EPA has not
allocated the resources needed to identify the scope of the problem and
to develop a strategy to address issues related to contaminated sediments. To develop an effective long-term contaminated sediment strategy it is critical to know how much work remains to be done. To address
that question, one needs to have three pieces of information:
1. How much work remains at sites already categorized as contaminated sediment megasites.
2. How many contaminated sediment sites already on the NPL are
likely to be determined to be megasites.
3. How many new such sites are likely to be added to the NPL in
the coming years.
None of that information is readily available from EPA. Clearly,
EPA should not stop and wait until this information is collected. However, it is important that EPA obtain this information and update it regularly in order to be able to forecast likely future costs and needed resources, as well as to assess what kinds of research and monitoring
improvements are likely to have the largest benefit to the program.
Cleanup Under Authorities Other Than Superfund
Remediation of contaminated sediments is also conducted under
authorities other than Superfund and can be led by various parties, such
as state or federal agencies or private entities, in combination or individually. For example, a 5-mile reach of the Grand Calumet River, a
highly industrialized tributary to Lake Michigan in northwest Indiana,
was dredged by U.S. Steel Corporation pursuant to a Clean Water Act
33
consent decree and a Resource Conservation and Recovery Act corrective-action consent order (Menozzi et al. 2003). This project, described as
“the largest environmental dredging project to be undertaken in North
America,” removed 786,000 cy of sediment from the Grand Calumet
River (U.S. Steel 2004).
State programs conduct and oversee sediment remediation under a
variety of authorities. For example, the State of Washington Department
of Ecology is charged with cleaning up and restoring contaminated sites
under authority of the Model Toxics Control Act (MTCA) and Sediment
Management Standards (SMS) (Washington Department of Ecology
2005). In 2005, 142 sediment cleanup sites were identified in Washington:
41 were being cleaned up under federal authorities, 48 were using state
authority alone, 11 were under federal and state authorities, and the remaining 42 were either voluntary (conducted by the responsible party)
or the authority had not been assigned (Washington Department of
Ecology 2005).
Contaminated sediments in many harbors and rivers of the Great
Lakes are addressed in the Great Lakes Water Quality Agreement between the United States and Canada, which established 43 areas of concern (AOCs) in U.S. and Canadian waters. The U.S. EPA Great Lakes
National Program Office administers funds from the Great Lakes Legacy
Act of 2002 for the remediation of contaminated sediment at AOCs (EPA
2004c). The first Legacy Act cleanup was in 2005 at the Black Lagoon in
the Detroit River AOC near Trenton, MI. At that site, 115,000 cy of contaminated material was dredged, and the area was capped. Hog Island,
near Superior, WI, in the St. Louis River AOC of Lake Superior, was
remediated with dry excavation (see Sediment Management Techniques
in this chapter for a description of remedial methods). In 2006, two projects were under way with Great Lake Legacy Act funds. The Ruddiman
Creek remedial action in Muskegon, MI, contains an excavation and
dredging component and is expected to remove around 80,000 cy.
Dredging will also occur at the Ashtabula River, near Cleveland, OH,
where it is expected that about 600,000 cy of contaminated sediment will
be removed from the lower portion of the river.
Another program, the Urban Rivers Restoration Initiative, is a collaboration between EPA and the U.S. Army Corps of Engineers for urban-river cleanup and restoration (EPA 2003a). Eight demonstration pilot projects, including a dredging project in the Passaic River in New
34
Jersey, have been developed to coordinate the planning and implementation of projects to promote clean water and sediment among multiple
jurisdictions and federal authorities.
This section is by no means a comprehensive listing of sediment
projects or efforts outside of Superfund; rather, the intent is to convey
that there are many sediment remediation projects outside of Superfund
that are conducted by multiple groups and under several authorities. To
the extent that other environmental dredging activities are conducted to
address risk from contaminated sediments, many of the discussions and
conclusions presented in the latter chapters of this report will be applicable.
EVALUATING RISK REDUCTION AT CONTAMINATED
SEDIMENT SITES
Risks Posed by Contaminated Sediment
As briefly described in Chapter 1, contaminants in sediment can
pose risks to human health and the environment. Apart from direct exposure to contaminated sediment during, for example, recreational activities, humans typically are exposed to contaminants through the ingestion of fish or wildlife that have accumulated contaminants from the
sediment. Fish and wildlife are exposed to contaminants in sediments
through a number of pathways, including absorption from pore water or
sediments, incidental ingestion of contaminated sediments, and consumption of contaminated organisms. Several of those processes are presented graphically in Figure 2-1. Predicting effects of exposures can be
complex. Variations in the sediment environments will alter the
bioavailability of contaminants, and this can markedly affect their accumulation and effects on organisms (NRC 2003). For instance, the presence of sulfide will greatly decrease the bioavailability of many metals,
and organic carbon can decrease the bioavailability of organic pollutants,
such as PCBs (EPA 2003b, 2005c).
Accessibility is a primary factor in exposure to and effects of contaminated sediments. A common problem in assessing risks posed by
contaminated sediments is that the contaminants (or the highest contam-
35
FIGURE 2-1 Generic conceptual site model indicating contaminated sediment
exposure pathways between sediment and ecologic receptors, including fish,
shellfish, benthic invertebrates, birds, and mammals. Source: EPA 2005a.
inant concentrations) can be buried beneath relatively clean sediment
that has deposited over time (see Figure 2-2).
Because sediment contaminants typically are strongly associated
with the sediment particles, contaminants buried below the biologically
active zone are neither accessible nor available to sediment- or waterdwelling organisms. In such cases, a relatively small continuing source
may pose a greater risk of exposure and associated injury than a large
buried inventory of sediment associated contaminants. Risk due to
sediments is usually limited to contaminants that are present in or can
migrate into the biologically active zone, the upper layers of sediment
where organisms live or interact. That layer typically ranges from a few
centimeters to 10-15 cm deep, although some organisms (including
aquatic plants) may penetrate more deeply (Thoms et al. 1995; NRC
2001).
Sediment-associated contaminants tend to collect in relatively stable depositional zones in water bodies. In such environments, buried
contaminants (that is, those below 10-15 cm) may never be exposed to
the biologically active zone. However, water bodies are dynamic systems
and even in generally depositional and stable environments, high flow
events and changes in hydrologic conditions may lead to short term
)
36
FIGURE 2-2 Historical changes in sediment core profiles of mercury concentrations collected from Bellingham Bay, WA, during natural recovery (dredging or
capping was not performed). Sediment cores were taken over time (1970, 1975,
1996) from the same vicinity of Bellingham Bay in an area with a stable sediment
deposit. At this site, mercury was released from a nearby facility from 1965 until
controls were put in place in 1971. Sedimentation has since continued to bury the
contaminants. Source: Patmont et al. 2004. Reprinted with permission from the
authors; copyright 2003, Battelle Press.
erosion, exposure, and transport of these contaminants to the biologic
active zone. In environments subject to such conditions, removal of contaminant mass may be an effective remedial response to the risk posed
by them. If contaminants buried below the biologically active zone are
likely to remain buried, the potential exposure and risk may be so small
that remediation of any kind is unwarranted. Remediation of deeply buried contaminated sediments that do not contribute to the exposure of
aquatic systems now or under future conditions will not achieve risk
reduction goals. In such cases, other contaminant sources, for example
inadequately controlled surface discharges or atmospheric deposition,
may control exposure and risk. A fair amount of effort in recent years
has gone into developing approaches for assessing sediment column stability and refining hydrodynamic models and linking them with fate and
transport models to estimate contaminant transport under various condi-
37
tions (e.g., Bohlen and Erickson 2006). Output from these approaches
and models are important in estimating the risks associated with remedial alternatives.
Decision-Making in a Risk-Based Framework
Principles for understanding and comparing risk reduction from
various sediment remediation techniques are discussed briefly below,
however, it should be noted that it is not the mandate or intent of the
report to develop specific recommendations and procedures for performing comparative risk analyses of remedial alternatives in selection of a
sediment remedy. While important, that type of detailed assessment was
not requested or undertaken. The brief discussion provided here on riskbased remedy selection is intended to provide background for later discussions on improving decision making.
The process of managing risk at contaminated sediment sites was
evaluated extensively in the 2001 National Research Council report A
Risk Management Strategy for PCB-Contaminated Sediments (NRC 2001).
Perhaps its most relevant conclusion is that all decisions regarding the
management of PCB-contaminated sediments should be made in a riskbased framework. The report further suggests that the framework developed by the Presidential/Congressional Commission on Risk Assessment
and Risk Management provides a good foundation for assessing the
risks and the management options for a site (see Box 2-3). The general
framework exhibits several key features that make it appropriate for the
management of contaminated sediment sites. It recognizes that riskreduction should be the foundation of any decision-making process and
the importance of the participation of interested and affected stakeholders in the decision-making process. It also provides a systematic and
structured process for identifying and assessing risks, evaluating and
implementing options, and monitoring the success of the overall process.
The 2001 Research Council report also recommends that risk assessments
and risk management decisions be site specific and concluded that current management options can reduce risks but cannot eliminate PCBs
and PCB exposure from contaminated sediment sites. Because all remedial options will leave residual PCBs, the short- and long-term risks that
38
they pose should be considered in evaluating management strategies.
Those ideas also apply to other sediment contaminants.
BOX 2-3 Presidential/Congressional Commission on Risk
Assessment and Risk Management
The Presidential/Congressional Commission on Risk Assessment and Risk
Management was formed in response to the 1990 Clean Air Act Amendments in
which Congress mandated that a Commission on Risk Assessment and Risk
Management be formed to “make a full investigation of the policy implications
and appropriate uses of risk assessment and risk management in regulatory programs under various Federal laws to prevent cancer and other chronic human
health effects which may result from exposure to hazardous substances”
(PCCRARM 1997, p. i).
The commission ultimately developed a report that introduced a risk management framework “to guide investments of valuable public sector and private
sector resources in researching, assessing, characterizing, and reducing risk”
(PCCRARM 1997, p. i). The commission proposed a six-stage process:
•
•
•
•
•
•
Define the problem and establish risk management goals.
Assess risks associated with the problem.
Evaluate remediation options for addressing the risks.
Select a risk management strategy.
Implement the risk management strategy.
Evaluate the success of the risk management strategy.
This process should be conducted
• In collaboration with all affected parties.
• In an iterative fashion when substantive new information becomes
available.
The proposed process, depicted in the schematic below, is a systematic
method to manage risks that the commission defined as “the process of identifying, evaluating, selecting, and implementing actions to reduce risk to human
health and to ecosystems. The goal of risk management is scientifically sound,
cost-effective, integrated actions that reduce or prevent risks while taking into
account social, cultural, ethical, political, and legal considerations” (PCCRARM
1997, p. 2).
39
Key features of the framework are recognition of the importance of stakeholders in the process, the importance of defining risks in a broader context than
single risks associated with single chemicals in single environmental media, and
the importance of an iterative process where earlier decisions can be revisited
when new findings are made.
Remedy selection is a complex process with many considerations
(see Box 2-1). In some cases, removal will be the best option for risk reduction and satisfying the NCP criteria, in others, capping or monitored
natural recovery will be preferable. An analysis of alternative remedies
typically includes a comparison of both the short- and long-term risks to
human and environmental receptors associated with a particular site. For
example, the risks from dredging can include exposure to contaminants
during dredging, rehandling, and transport, and contaminants that remain after operations are completed. Those risks would be compared to
other alternatives, including risks from unconfined contaminated sediment and potential future resuspension and transport during storm and
non-storm events. Risks beyond those related directly to exposure to
contaminants are also considered in this process (see Net Risk Reduction
below).
Technical and policy guidance for making remedial decisions using
a risk based framework at contaminated sediment sites was recently issued (EPA 2005a). The document provides a useful evaluation of the
various sediment management approaches and their advantages and
limitations in attaining risk reduction. It discusses in detail aspects of the
40
Superfund decision-making process (site characterization, feasibility
study, and remedy selection) particular to contaminated sediment, and it
offers recommendations for implementing an effective monitoring plan.
The guidance concludes that “The focus of remedy selection should be
on selecting the alternative best representing the overall risk reduction
strategy for the site according to the NCP nine remedy selection criteria…. EPA’s policy has been and continues to be that there is no presumptive remedy for any contaminated sediment site, regardless of the
contaminant or level of risk” (EPA 2005a, p. 7-16).
Measuring Risk-Reduction
Estimating the degree of risk reduction is central in considering the
potential effectiveness of a remedial action. Risk posed by chemical contamination is a function of the duration and intensity of exposure and
the ability of the chemical or chemical mixture to exert adverse effects.
There is not a direct measure of risk, so surrogate metrics are used to
estimate risk. Environmental analyses have to use metrics that, in practice, can be employed relatively easily, are not time- and cost-prohibitive,
and have sufficient accuracy and precision to be reliable. Estimates of
risk reduction at contaminated sites have often centered on measuring
changes in the mass, volume, and concentration of contaminated sediments. Those measures are related to the potential for exposure in the
aquatic environment but do not provide information on effects. Therefore, although they are the most prevalent, they are not fully adequate to
describe risk or to chart risk reduction. Toxicity testing, biologic community indexes, and tissue-residue analyses provide a fuller picture of effects, although they too have limitations in their ability to describe risk.
(See Chapter 5 for further discussion on the metrics and their advantages
and disadvantages in estimating risk to aquatic biota and humans.)
Temporal Scale
In characterizing risk and evaluating risk reduction, it is necessary
to consider the duration of time over which exposure and effects occur.
After remediation, risk is usually predicted to decline over time rather
than reach a protective level immediately on completion of the remedy,
41
so remedy selection involves a comparison of time profiles of predicted
risks, often on scales of decades. Such a comparison of risk profiles over
time is how the long-term effectiveness of dredging is evaluated relative
to alternative technologies for the largest and most complex sites. Factors
dictating the time to reduce risk are site specific and include the time
required to design and fully implement the remedy, the time required to
cleanse the food chain of existing contaminant body burdens, and the
time for natural recovery processes to attenuate any residual surface
sediment concentrations after implementation is complete.
Net Risk Reduction
The 2001 National Research Council report indicated that the
paramount consideration for contaminated sites should be the management of overall or net risks to humans and the environment in addition
to specific risks. The report concludes that the evaluation of sediment
management and remediation options should take into account all costs
and potential changes in risks for the entire sequence of activities and
technologies that constitute each management option. (For example,
managing risks from contaminated sediments in aquatic environments
might result in the creation of additional risks in both aquatic and terrestrial environments.) The report also suggests that a broader array of
risks—including societal, cultural, and economic risks—should be evaluated comprehensively. The concept of net risk reduction has been embraced by EPA in its Contaminated Sediment Remediation Guidance for Hazardous Waste Sites (EPA 2005a); it states that “Project managers are
encouraged to use the concept of comparing net risk reduction between
alternatives as part of their decision-making process for contaminated
sediment sites, within the overall framework of the NCP remedy selection criteria. Consideration should be given not only to risk reduction
associated with reduced human and ecologic exposure to contaminants,
but also to risks introduced by implementing the alternatives. The magnitude of implementation risks associated with each alternative generally
is extremely site specific, as is the time frame over which these risks may
apply to the site. Evaluation of both implementation risk and residual
risk are existing important parts of the NCP remedy selection process”
(EPA 2005a, pp. 7-13, 7-14).
42
Risk-Based Objectives and Cleanup Levels
Each site has its own set of contaminants with different concentrations and distributions in its own particular geologic, geochemical, geographic, social, ecologic, and economic setting. Therefore, management
decisions based on the above framework are expected to differ among
sites.
At Superfund sites, the overall goal of sediment management is reduction of risk to human health and the environment. That goal takes the
form of remedial action objectives, which are used in developing and
comparing alternatives for a site, and typically describe the desired effect
of the remediation on risk (for example, reduction to acceptable levels of
the risks to people ingesting contaminated fish). Attainment of remedial
action objectives can be difficult to quantify or might occur in a time
frame or encompass a spatial scale that makes it difficult to link to remedial actions. Under such circumstances, cleanup levels, such as achievement of a sediment concentration or removal of a given mass of contaminant or sediment, which can be more easily used to evaluate
remedial actions, are often adopted (EPA 2005a). Ideally, clean-up levels
are tied to effects-based risk thresholds, and take into account effects of
combinations of contaminants.
That the application of a good risk management strategy is likely to
result in significantly different cleanup levels at different sites makes it
difficult for the committee to draw conclusions about the expected effectiveness of dredging. At some sites, cleanup levels are far less stringent
than at others, and thus all other things being equal, a site with less
stringent cleanup goals is more likely to be “successful” than a site with
more stringent goals. Geologic and site-specific conditions also differ, so
even if the cleanup goals are similar at different sites, the technical ability
to reach the goals may differ. Thus, one needs to be highly cautious in
suggesting that success with a remedial option at one site necessarily
means that the same success is likely at another.
To help to ensure that remedial actions achieve their desired objectives, 11 principles have been developed by EPA to guide sediment
remediation (Box 2-4). The principles were developed partially in response to the recommendations of the National Research Council Committee on Remediation of PCB-Contaminated Sediments (NRC 2001).
43
BOX 2-4 Eleven Principles of Contaminated Sediment Management
In February 2002, the EPA Office of Solid Waste and Emergency Response
promulgated 11 principles of contaminated sediment management (OSWER
Directive 9285.6-08):
1. Control sources early.
2. Involve the community early and often.
3. Coordinate with states, local governments, tribes, and natural-resources
trustees.
4. Develop and refine a conceptual site model that considers sediment
stability.
5. Use an iterative approach in a risk-based framework.
6. Carefully evaluate the assumptions and uncertainties associated with
site-characterization data and site models.
7. Select site-specific, project-specific, and sediment-specific risk-management approaches that will achieve risk-based goals.
8. Ensure that sediment cleanup levels are clearly tied to risk-management
goals.
9. Maximize the effectiveness of institutional controls and recognize their
limitations.
10. Design remedies to minimize short-term risks while achieving longterm protection.
11. Monitor during and after sediment remediation to assess and document
remedy effectiveness.
The principles were designed to help EPA site managers to make scientifically sound and nationally consistent risk-management decisions at contaminated sediment sites. The principles are consistent with the recommendations of
the National Research Council A Risk Management Strategy for PCB-Contaminated
Sediments and the Presidential/Congressional Commission on Risk Assessment
and Risk Management. They were incorporated into the Contaminated Sediment
Remediation Guidance for Hazardous Waste Sites (EPA 2005a).
The Conceptual Site Model: A Working Understanding of Processes
Leading to Risk from Sediment Contamination
The development of remedial action objectives and cleanup levels
to reduce risk is based on a conceptual understanding of cause-effect
44
relationships among contaminant sources, transport mechanisms, exposure pathways, human receptors, and ecologic receptors at each affected
level of the food chain. That understanding of causal relationships is
known as a conceptual site model (CSM) (EPA 2005a) and is typically
developed for each site on the basis of site-specific conditions. The link
between risk and the inventory of contaminants in sediments is not always obvious. An accurate CSM is critical for identifying the processes
and pathways that might lead to risk and appropriate means of intervening to reduce risk. For example, evaluation of the stability or potential
instability of buried deposits and their potential for exposure in the
bioavailable zone is a key component of a CSM because a CSM must be
able to differentiate between important and unimportant routes of exposure.
In addition to linking site contaminant sources to exposures and
risks, the CSM must account for background conditions, including contaminant distribution from offsite sources. Ecosystems may be highly
stressed because of multiple watershed and atmospheric effects on conventional water quality measures, such as nutrients, suspended solids,
acidity, dissolved oxygen, and temperature. The contribution of background stressors or other background sources to site effects should be
evaluated, including assessment of their importance relative to the contaminants of concern, and recognized as potentially complicating factors
in ecosystem restoration.
The CSM should guide site investigation, and its hypotheses and
assumptions should be tested and refined as site data are acquired.
When the CSM has been accepted with a high degree of confidence, it is
used to define remedial action objectives. Basing remedial action
objectives on the best scientific understanding of the mechanisms that
lead to site-specific risk maximizes the likelihood that remedial actions
will meet the objectives.
For most sites, but especially for the largest and most complex, a
quantitative dimension must be added to the CSM to support development and selection of a remedy. The result is a mathematical model or a
set of models of the various component processes. Mathematical models
are used to quantify the same cause-effect relationships that are embodied in CSMs so that magnitudes of predicted outcomes can be associated
with specific causes or actions, such as contaminant loads, environmental conditions, and remedies. To be most accurate, the models
45
should be supported by and calibrated to site-specific data on the environmental media (such as sediments, pore water, and water), receptors
(such as benthic organisms, fish, and humans), and processes (such as
toxicity, bioavailabity, and bioaccumulation) that are being examined.
Mathematical models range from simple to complex, including analytic equations representing established scientific relationships between
independent and dependent variables, statistical cause-effect relationships between site variables, and systems of differential equations representing multiple fate and transport processes. With a mathematical
model, quantitative versions of hypotheses can be tested and refined on
the basis of site data, including data from field surveys of site conditions
and pilot studies of remedial technologies, and then the relative effectiveness of alternative remedies in reducing exposures can be estimated,
including the sensitivity of exposure to the remedies and the time
needed to reduce exposure. The measures of predicted effectiveness are
used to support remedy selection.
Models are subject to uncertainty because of the uncertainty in parameters and process representations. Model testing and refinement
does not end with the selection of a remedy through a record of decision.
It is important that the conceptual and mathematical site models also be
used in designing monitoring of conditions during implementation and
post-remedy phases and that the monitoring data be used to validate the
models’ predictions. When risk reduction deviates significantly from a
model’s predictions, the model should be modified or recalibrated to
improve its accuracy so that more reliable predictions can be available to
guide midcourse adjustments in the remedy (EPA 2005a).
CONTAMINATED SEDIMENT MANAGEMENT TECHNIQUES
Contaminated sediment is managed with various techniques, including source control, natural recovery, capping, and removal (dry excavation and dredging). Removal necessitates management of the removed material, which normally includes dewatering, transport, and
disposal. Treatment of dredged material to remove or destroy contaminants is an option, but cost and other factors usually lead to disposal in
upland landfills or in near-shore confined disposal facilities. In some
46
cases, dredged material can be returned to the aquatic environment
through containment in confined aquatic disposal facilities (EPA 2005a).
The National Research Council Committee on Contaminated Marine Sediments (NRC 1997) and Committee on Remediation of PCBContaminated Sediments (NRC 2001) have reviewed and reported on a
number of sediment management techniques. The committees stated that
source control is advisable in all contaminated sediment management
projects, notwithstanding the difficulties of identifying some sources of
contamination. Beyond source control, interim controls (temporary
measures to address exposures immediately) and long-term controls
(such as in situ management technologies, sediment removal and transport, and ex situ management) may be needed to address sediment contamination.
More recently, EPA (2005a) lists both in situ and ex situ remedial
strategies for managing risks posed by contaminated sediment. The in
situ strategies include monitored natural recovery (MNR), in situ capping, hybrid (thin-layer placement) approaches, institutional controls,
and in situ treatment; the ex situ strategies include dredging and dry
excavation (following dewatering or water diversion). See Box 2-5 for an
explanation of these approaches. The present committee’s focus is on
environmental dredging, which is conducted specifically to remove contaminated sediments, as opposed to navigational dredging, which typically is intended to maintain depth in waterways for navigation or other
purposes.
HISTORICAL PERSPECTIVE ON THE USE OF
REMOVAL TECHNOLOGIES TO REDUCE RISK
Although there has never been a presumptive remedy for sediments, the historical preference for removal is evident in the large percentage of sites whose remedy was based entirely or in part on dredging.
In an overview of Superfund sediment remediation, EPA presented information from 60 tier 1 sites for which a remedy had been selected. Of
the 60, 57% had only removal as the remedial action, 15% capping with
removal, 13% removal with MNR, 5% only capping, 2% only MNR, and
8% all three remedies (Southerland 2006). The historical preference for
47
BOX 2-5 Remedial Approaches to Contaminated Sediment
In Situ Approaches
In Situ Approaches
•
Monitored natural recovery (MNR) is a remedy for contaminated sediment
that typically relies on naturally occurring processes to contain, destroy, or reduce the
bioavailability or toxicity of contaminants in sediment.
•
In situ capping refers to the placement of a subaqueous covering or cap of
clean material over contaminated sediment, which remains in place.
•
Hybrid approaches refers to placement of a thin layer of sand or other material to accelerate recovery.
•
Institutional controls are controls on the use of resources. They typically include fish-consumption advisories, commercial fishing bans, waterway-use or landuse restrictions (for example, no-anchor or no-wake zones and limitations on navigational dredging), and agreements on maintenance of dams or other structures.
Ex Situ Approaches
•
Dredging and excavation are common means of removing contaminated bottom sediment from a body of water either while the sediment is submerged (dredging) or after water has been diverted or drained (excavation). Ex situ approaches can
include backfilling with clean material as needed or appropriate.
removal is probably based on the perception (in both agencies and the
public) of the permanence of the remedy. Dredging and excavation remove the mass of contaminants from the aquatic environment, and this
has historically been viewed as key to reducing human health and environmental risks.
Technologies for removing sediment were already well established
in the early years of sediment cleanup, in part as an extension of remediation technologies applied at upland sites. Most of the initial technologies for managing sediment came from the U.S. Army Corps of Engineers experience with navigational dredging and disposal. Other
remedies were typically viewed as less certain by regulators and the
public with respect to long-term effectiveness or permanence. Leaving
contamination in place under a capping or MNR remedy was often considered more uncertain because of the residual risk posed by contaminants left in place.
48
The dynamic nature of aquatic environments has often led to the
selection of removal as the preferred alternative in many areas of the
country. Contaminated sediment is often associated with industrial, urban harbors where operational and navigational constraints are viewed
as limiting the feasibility of capping or natural recovery. Those environments are often subject to disturbances, such as those caused by prop
wash, seasonal flooding, ice scour, and storm surges, which were viewed
as creating substantial risk if contaminants were left in place.
Removal of contaminated sediment has brought unique challenges
that were initially not well recognized. Navigational dredging techniques adopted for environmental dredging are designed to achieve a
specific bottom elevation or the removal of a specific volume, often in the
shortest possible time, whereas environmental dredging typically must
achieve a specific final concentration while minimizing contaminant releases during dredging, handling, and disposal. As dredging remedies
have been implemented at various sites, the effects of resuspension and
transport of contaminated material off site and residual contamination in
a remediated area have become apparent (Bridges et al. in press). The
risks associated with the implementation of environmental dredging
have received a great deal of attention in the last few years (EPA 2005a;
Wenning et al. 2006).
Greater experience with capping remedies has been gained over the
last decade; cap performance can be better predicted and quantified, and
this has led to greater acceptance among agencies. In addition, capping
typically has been less expensive and can be implemented more quickly,
so it is often preferred by responsible parties (Palermo et al. 1998). In response to the increasing experience with remedial technologies, recent
guidance from EPA has called for a more equitable evaluation of all
remedies with careful analysis of the short-term and long-term risks associated with any remedy and thorough consideration of site-specific
conditions (EPA 2005a).
OVERVIEW OF ENVIRONMENTAL DREDGING
Dredging refers to the removal of sediment from an underwater environment. It involves dislodging and removing material on the bottom
of a waterway. Dredges are normally classified according to the basic
49
operation by which sediment is removed, such as mechanical or hydraulic10 (EPA 1994). For purposes of this report, excavation in the dry using
conventional equipment operating within dewatered containments such
as sheet-pile enclosures or cofferdams is not covered. The term environmental dredging is more generally associated with removal of sediment
from under water. Environmental dredging can be accompanied by
backfilling of the dredged areas. Placement of clean material covers and
mixes with dredging residuals and further reduces risk from contamination that remains after dredging. Unlike capping, permanent confinement of underlying material is usually not the goal.
Typical objectives of environmental dredging are shown in Box 2-6.
Because the purpose of navigational dredging is to restore navigable
depth to a waterway, the selection of equipment and operational approaches considers economics, effectiveness, and environmental protection (USACE/EPA 1992) in that order. Conversely, environmental dredging has remediation as its stated purpose. The distinction results in
reversing the order of importance of the selection factors for equipment
and operational approaches; that is, one needs to consider environmental
protection and effectiveness first before considering economics (Palermo
et al. 2006).
BOX 2-6 Objectives of Environmental Dredging
• Dredge with sufficient accuracy such that contaminated sediment is
removed and cleanup levels are met without unnecessary removal of clean
sediment.
• Dredge the sediments in a reasonable period of time and in a condition
compatible with subsequent transport for treatment or disposal.
• Minimize and/or control resuspension of contaminated sediments,
downstream transport of resuspended sediments, and releases of contaminants
of concern to water and air.
• Dredge the sediments such that generation of residual contaminated
sediment is minimized or controlled.
Source: Palermo et al. 2006.
Pneumatic systems, which use compressed air to pump sediment out of a
waterway, have not gained general acceptance in environmental dredging projects in the United States.
10
50
Types of Environmental Dredges
Selection of dredging equipment is sediment specific, site specific,
and operations specific. Many textbooks and manuals describe the science and engineering principles of dredges, their selection, and their operation (Bray 1979; USACE 1983; Herbich 2000). This section provides
basic definitions of dredging methods and equipment types normally
considered for environmental dredging. There is no attempt to list all the
possible types of dredge equipment that may be applicable to environmental dredging. Box 2-7 lists the equipment most commonly used for
environmental dredging according to type (category) and definition
(Palermo et al. 2004). Figure 2-3 shows the basic dredge types. More detailed descriptions of environmental-dredging equipment are available
elsewhere (Averett et al. 1990; EPA 1994; EPA 2005a).
Other dredge types—such as hopper dredges, dustpan dredges,
and bucket-ladder dredges—are not included in Box 2-7, because they
are used primarily for navigational dredging. In addition, within dredge
types, specific designs may differ and may have varied capability. In
general, the dredge types listed above represent equipment that is readily available and used for environmental dredging projects in the United
States.
A number of newer dredges, including some specifically designed
for environmental dredging, are available. They have been termed specialty dredges and are intended to provide benefits by reducing sediment resuspension and contaminant releases. Other advantages may
include operational efficiency for removal of sediment and transportation, depending on the sediment and project conditions and the performance standards. Most specialty dredge designs originated outside
the United States, but several U.S. companies have now formed partnerships that allow use of specialty equipment from various countries. Field
experience with specialty dredges in the United States is limited (Palermo et al. 2003). The dredges have been proposed for use at contaminated sediment sites, but little information is available about their sediment-extraction efficiency or about the claimed improvements in
innovations, such as improved solids capture and reduced resuspension.
The equipment used for environmental dredging is usually smaller
than that commonly used for navigation dredging because removal
51
BOX 2-7 Equipment Commonly Used in Environmental Dredging
Mechanical Dredges
• Clamshell: Wire-supported conventional open clam bucket.
• Enclosed bucket: Wire-supported, nearly watertight or sealed bucket. In
contrast to conventional open buckets, recent designs also incorporate a level-cut
capability instead of the circular cut of conventional buckets (for example, the
horizontal profiling buckets).
• Articulated mechanical: Backhoe design, clam-type enclosed bucket,
hydraulic closing mechanism, all supported by articulated fixed arm.
Hydraulic Dredges
• Cutterhead: Conventional hydraulic pipeline dredge with conventional
cutterhead.
• Horizontal auger: Hydraulic pipeline dredge with horizontal auger
dredgehead.
• Plain suction: Hydraulic pipeline dredge using a dredgehead design
with no cutting action and plain suction (for example, cutterhead dredge with no
cutter basket mounted, matchbox dredgehead, and articulated scoops).
Pneumatic Dredges
• Pneumatic: Air-operated submersible pump, pipeline transport, and
wire-supported or fixed-arm-supported.
Specialty Dredges and Diver-Assisted Dredges
• Specialty dredgeheads: Other hydraulic pipeline dredges with specialty
dredgeheads or pumping systems.
• Diver-assisted: Hand-held hydraulic suction with pipeline transport.
Source: Adapted from Palermo et al. 2004.
volumes and rates tend to be lower and water to be shallower. Mechanical-bucket sizes range from 2 to 8 m3 (about 3 to 10 cy), and hydraulicpump sizes range from 15 to 30 cm (about 6 to 12 in.) (Palermo et al.
2006). Obviously, larger dredges are available for both mechanical and
hydraulic equipment and can be used for environmental dredging if
needed.
52
FIGURE 2-3 Categories of dredging and sediment removal equipment. Source:
Francingues and Palermo 2006.
Dredging—One Part of the Overall Process Train
Physically removing sediments by dredging is only one component
of the overall remediation process. The key processing steps shown in
Figure 2-4 include (EPA 2005a):
•
Mobilization and setup of equipment.
53
• Site preparation including debris removal and protection of
structures.
• Removal (environmental dredging).
• Staging, transport, and storage (rehandling).
• Treatment (pretreatment, solidification and stabilization of solids, treatment of decant water and/or dewatering effluents and sediment, and potentially separate handling and treatment of materials with
and without special requirements under the Toxic Substances Control
Act [TSCA]).
• Disposal (liquids and solids).
Environmental dredging must be compatible with all later steps in
the process train. For example, the production rate of a dredge (either
mechanical or hydraulic) depends heavily on the mode of transportation
and the ability to rehandle or directly manage the dredged material on
the other end of the process. Compatibility must be considered with respect to the type of pretreatment, treatment, and disposal being planned,
especially the availability, size, and capacity of disposal sites, the distance from dredging site to treatment or disposal sites, and constraints
associated with production rates for transport, storage, rehandling,
treatment, or disposal. Inefficiencies in remedial dredging projects can
result from constraints associated with components of the remedy other
than dredging, such as dewatering capacity, water-treatment effectiveness, and disposal location and capacity (Palermo et al. 2006).
Dredging accounts for only part of the overall cost of an environmental-dredging project. In a complex project, large costs may be associated with the transport, dewatering, and ultimate disposition of the
dredged material. Recent data, described below, support the premise
that dredging accounts for 10-20% of the total cost of an environmental
dredging project. For example, EPA Region 1 (EPA 2005d) reported on
the costs of the 2004 New Bedford Harbor Superfund dredging project,
Mobilization
Mobilization/
Setup
andSetup
Removal
Site
(Dredging or
Preparation
Excavation)
Removal
(Dredging or
Excavation)
Staging
and
Staging/
Transport
and
Transport/
Rehandling
Re-handling
Treatment
Disposal
FIGURE 2-4 Dredging-process train. Source: Adapted from Palermo et al. 2006.
54
as shown in Figure 2-5; dredging itself represents only 17% of the total
yearly construction and operations cost. Similarly, in the Head of Hylebos remediation (Figure 2-6) dredging operations conducted from 2003
to 2006 (Dalton, Olmsted & Fuglevand, Inc. 2006), dredging represents
17% of the total cost. Dredging cost varies widely, depending on many
factors, including site conditions, the nature of the sediments and contaminants, the type and size of dredge selected, production rates, and
seasonal construction windows. However, when dredging is selected as
a remedy, all the other components of the process train will probably be
required and will account for most of the overall cost. Only in those cases
where transportation and disposal of sediments are relatively inexpensive (for example, where there is an existing in-water or upland disposal
site, both suitable for the long-term containment of contaminated sediment and in close proximity to the dredging) will the dredging be a major cost element.
TECHNICAL ISSUES ASSOCIATED WITH DREDGING
Environmental dredging typically strives to achieve contaminantspecific cleanup levels set at each site. A number of technical issues can
limit ability or efficiency in achieving those levels. This section describes
several of the issues, many of which are revisited in the context of site
evaluations in Chapter 4.
Accuracy of Dredging vs. Accuracy of Sediment Characterization
The benefits of being able to position a dredge cut accurately may
be achieved only if a corresponding degree of accuracy is reflected in the
site and sediment-characterization data. The ability to map the precise
location of chemical concentrations accurately both horizontally and vertically depends on the data density (grid density), accessibility of deeper
sediments, and other aspects of site characterization (Palermo et al.
2006). In some cases, the ability to locate the dredge cut accurately exceeds the accuracy of the knowledge of the location of the contaminated
sediments (Palermo et al. 2006).
Monitoring P M &
Fees 25%
55
Dredging 17%
Desanding 12%
T ransport &
Disposal 22%
Dewatering
18%
Water T reatment
6%
FIGURE 2-5 Cost breakdown of components of environmental dredging at the
New Bedford Harbor (New Bedford, MA) project in 2004. Full-scale operations
cost about $800,000 per week. PM = project management. Source: Data from EPA
2005d.
Monitoring PM
& Fees
15%
Other
4%
Dredging
17%
Mob/Demob
15 %
T ransport &
Disposal
31%
Wat er Management
6%
Sediment Handling
12%
FIGURE 2-6 Cost breakdown of components of environmental dredging at the
head of the Hylebos Waterway (Commencement Bay, WA) project. PM = project
management; Mob/Demob = mobilization and demobilization. Source: Data from
P. Fuglevand, Dalton, Olmsted & Fuglevand, Inc., personal commun., May 11,
2007.
In the context of this discussion, vertical operating accuracy is the
ability to position the dredgehead at a desired depth or elevation for the
cut, whereas vertical precision is the ability to maintain or repeat the vertical position during dredging. The key to the success of an environmental-dredging project is the removal of the target layer, which is de-
56
lineated by the cut line, without unnecessary removal of clean material
(Palermo et al. 2006).
The ability to dredge to a specified cut line in the sediment has
been greatly improved by the advent of electronic positioning technologies, such as differential global positioning systems (DGPSs) and kinematic differential global positioning systems (KDGPSs). Depending on
site conditions, dredge operator ability, size and type of dredge, and positioning instrumentation and software, the dredgehead and cut elevation may be locatable with vertical accuracy of less than 30 cm. Vertical
accuracies of 10 cm for fixed-arm dredgeheads should be consistently
attainable, whereas vertical accuracies of 15 cm should be attainable with
proper operator training in the use of wire-supported buckets (Palermo
et al. 2006).
Notwithstanding the previous statements regarding accuracies and
positioning of dredging equipment, there are numerous challenges in
attaining them. For example,
• Effective use of sophisticated dredge positioning systems requires sophisticated operators and contractors in order to achieve the
stated accuracies.
• In order to get effective positioning with any of the software
packages, the operators must be specifically trained and capable of system operation, and the systems must be properly operated and calibrated.
• Experience has shown that some systems are more difficult to
operate than others, and some systems may experience difficulties maintaining calibration. Simply using an electronic positioning system on a
dredge does not guarantee that the stated accuracy will be achieved.
Resuspension, Residuals, and Release of Contamination
All dredging equipment disturbs sediment and resuspends some
fraction of it in the water column. Resuspended sediment and the associated contaminants can settle back to the bottom in the dredge cut; finergrained materials can remain in the water column and be transported to
other locations. Those materials are deposited as residuals and result
from dredging. Dissolved contaminants may also be released to the wa-
57
ter column during dredging from resuspended or exposed contaminated
sediment. Figure 2-7 is a conceptual illustration of environmental dredging and those processes.
Dredged sediment resuspension, release, and residual and the resulting risk (the “4 Rs”) were the focus of a recent workshop held at the
U.S. Army Engineer Research and Development Center in Vicksburg,
MS (Bridges et al. in press). Effective remediation by dredging requires
minimizing the 4 Rs while maximizing the fifth R, removal—either the
dredging production rate or the volume removed (Francingues and
Thompson 2006). The type and amount of sediment resuspension, contaminant release, and residuals during a dredging operation depend on
many site-specific project factors, as shown in Box 2-8.
Resuspension
Resuspension is the process by which dredging and associated operations result in the dislodgement of embedded sediment particles,
which disperse into the water column. Resuspended particles may settle
in the dredging area or be transported downstream. Recent EPA guidance for sediment remediation states that
When evaluating resuspension due to dredging, it generally is important to compare the degree of resuspension to the natural sediment resuspension that would continue to occur if the contaminated sediment was not dredged, and the length of time over which
increased dredging-related suspension would occur.… Some contaminant release and transport during dredging is inevitable and
should be factored into the alternatives evaluation and planned for
in the remedy design.… Generally, the project manager should assess all causes of resuspension and realistically predict likely contaminant releases during a dredging operation (EPA 2005a, pp. 621, 6-22).
Resuspension concerns related to dredging include the physical effects of turbidity and burial that can result in seasonal restrictions on
dredging operations (dredging windows). Sediment resuspension can
58
Release
(Air)
Release
(Water)
Resuspension
Residual
(Sediment)
Removal
Residual
FIGURE 2-7 Conceptual illustration of environmental dredging and processes.
Source: Patmont 2006. Reprinted with permission; copyright 2006, Anchor Environmental LLC.
result in chemical releases to the water column (for example, from pore
water displaced from the dredged sediment or by desorption from resuspended sediment particles) and residual contamination on the bottom
after dredging. Resuspension can be caused not only by dredging
equipment but by propwash of tenders (push boats or tugs used to move
equipment) and during rehandling and transport operations, such as
filling and overflowing of barges and leaky pipelines. Estimates of resuspension from environmental-dredging projects range up to 10% of
the mass of sediment dredged (Patmont 2006). Rates of resuspension
depend on equipment, material, operator, and other site-specific factors.
Residuals
Residuals are contaminated sediment that remains after dredging.
There are two general types of residuals: generated residuals, contaminated sediment that is dislodged or suspended during dredging and
later redeposited within or adjacent to the dredging footprint; and undisturbed residuals, contaminated sediments found at the post-dredge sedi-
59
BOX 2-8 Site-Specific Factors Affecting Resuspension, Release, and
Residuals Sediment Physical and Chemical Properties
Sediment Physical and Chemical Properties
•
Grain size distribution (for example, percentages of silt, clay, and sand).
•
Organic carbon content.
•
Amount of sulfides.
•
Spatial and vertical distributions of contaminants in the sediment (for example, layering).
Site Conditions
•
•
•
•
•
•
•
•
Water velocity and degree of mixing.
Water salinity, hardness, alkalinity, and temperature.
Type of substrate (for example, hardpan, bedrock or soft sediment).
Type and extent of debris in sediment.
Weather, such as storms that result in wind and waves.
Wakes from passing vessels.
Fluctuations in water elevation.
Depth and slope of area to be dredged.
Equipment
•
Type of dredge (for example, cutterhead pipeline, open or closed bucket,
and specialty dredgehead).
•
Methods of dredging.
•
Skill of operators.
•
Extent of tender-boat activity.
•
Methods of sediment transport and offloading.
Source: Adapted from Palermo et al. 2006.
ment surface that have been uncovered but not fully removed as a result
of the dredging operation (Bridges et al. in press).
Residuals may result from incomplete characterization, inaccuracies of dredging, mixing of targeted material with underlying materials
during dredging, fallback (dislodged sediment not picked up), and resettlement of resuspended sediments (Palermo et al. 2006). Also contributing to residual contamination are such processes as sloughing of sediment into the dredging cut and sloughing induced by bank or slope
60
failures. Site-specific factors, such as debris or limitation of dredging by
bedrock or hardpan can influence the amount of residuals. Box 2-9 describes specific processes during dredging that contribute to residual
formation.
The residual contaminant mass is typically limited to the upper few
inches of sediment, which is populated and actively processed by sediment-dwelling organisms (although in the case of undisturbed residuals
the depth can be substantially greater). That upper layer is subject to erosion and other physical and chemical processes that may promote release
into the overlying water because of the entrainment of water into the
dredged sediment, which causes physical (decreased consolidation) and
chemical (redox) changes in the residuals. Residual contamination may
also be attributable to sediment that was not dredged, because of the
dredger’s failure to meet dredge cutlines (either depth or areal targets) or
errors or incompleteness in site characterization that failed to identify
appropriate depth and areal extent of contaminated sediment.
Patmont (2006) compiled data on residuals from 12 environmentaldredging projects. Final generated residuals ranged from approximately
2 to 9% (average = 5%)11 of the mass of contaminant dredged during the
last production cut. There is little research on the amount of generated
residuals transported outside the dredge prism, but their presence has
been documented analytically (EcoChem Inc 2005) and visually with
sediment-profile imagery (Baron et al. 2005).
Release
Release is the process by which the dredging operation results in
the transfer of contaminants from sediment pore water and sediment
particles into the water column or air. Contaminants sorbed to resuspended particles may partition to the water column and be transported
downstream in dissolved form along with contaminants in the released
More recently, Patmont and Palermo (2007) analyzed a similar (though not
identical) dataset and found that final generated residuals ranged from approximately 2% to 9% (average = 4%) of the mass of contaminant dredged during the
last production cut.
11
61
BOX 2-9 Specific Processes Contributing to the Residual
Layer During Dredging
For mechanical dredging, processes that contribute the residual layer are
•
The erosion of sediment from around and within the bucket as it is
placed on the bottom, closed, and raised through the water column. The erosion
in the water column can be controlled with the use of enclosed buckets. However there can be significant resuspenion of contaminated sediment during the
closing of enclosed buckets, as the bucket vents expel sediment at high velocity.
•
The overflow of turbid water from the sediment haul barge, controlled
with restrictions on barge overflow and associated capture and treatment of the
turbid water.
For hydraulic dredging, processes that contribute the residual layer are
•
The spillage layer generated by hydraulic dredging associated with the
turning of the cutterhead or auger in the sediment. Hydraulic dredges are normally configured with the inlet of the suction pipe well above the lowest reach of
the rotating cutterhead or auger. That means that the mixed layer generated by
the cutterhead or auger is not fully removed by the suction pipe and consequently there is a “spillage layer” left behind after dredging.
•
Another source of residual sediment is resuspension by the rotating
cutterhead or auger, when sediment is displaced away from the cutterhead or
auger into the water column.
Dredging, either mechanical or hydraulic, can result in the formation of a residual layer through a variety of mechanisms including
•
The sloughing of the sidewalls and headwall of the dredge cut face
back on to previously dredged areas. This sloughing can be controlled through
the use of relatively thin dredge lifts (few feet each) and by including a final
cleanup pass of dredging once the bulk of sediment has first been removed
(“two pass dredge approach”). If not controlled, this bank sloughing can result
in a considerable residual layer forming on previously dredged areas.
•
The remolding of soft fine-grained sediment by the dredging process
can significantly reduce the strength of the material and generate a more liquid
like flowable residual layer in the dredging area. This flowable material can be
very difficult to capture with the dredge and result in a residual layer that is
62
BOX 2-9 Continued
difficult to manage and control once it is formed. The formation of this layer can
be reduced (not eliminated) by a controlled and precise removal program using
electronic, GPS-enabled dredge positioning and mechanical dredging. Once
formed, capture of the flowable layer can be accomplished with overdredging
into native substrate, provided that substrate is not hardpan or bedrock.
Sources: Adapted from Dalton, Olmsted & Fuglevand, Inc. 2006; Fuglevand and
Webb 2006, 2007; Hartman 2006.
pore water. Contaminants in the generated or undisturbed residuals may
be released to the water column by densification, diffusion and bioturbation (Bridges et al. in press).
Releases of contaminants from the aforementioned sources and
processes are considered to be up to about 5% of the contaminant mass
in the sediment dredged, but larger or smaller releases may be observed,
depending on site-specific factors and the type and operating characteristics of the dredge (Sanchez 2001; Sanchez et al 2002). The degree of contaminant release to the air and water is directly related to the degree of
sediment resuspension (and pore water release) and chemical properties
affecting the mass transfer of contaminants. Therefore, control of resuspension should have high priority at many dredging project sites that
involve contaminated sediment. Contamination can also be released
from sediment beds to the water column in soluble form without particle
resuspension (Thibodeaux and Bierman 2003; Erickson et al. 2005). That
suggests that the residual layer is also a contributor of contaminant release after dredging. Control of solids is important but is not always sufficient to prevent contaminant losses.
Impact on Risk
Risk can result from contaminant exposures driven by resuspension, production of residuals, and contaminant release. Those processes
are important because they can alter the accessibility bioavailability of
contaminants, create additional contaminant exposure pathways that
63
potentially affect the risk resulting from dredging, and may continue to
influence risk after remedial operations cease. Surface-water concentrations and surface-sediment concentrations may increase during and after
dredging and can result in adverse effects and accumulation of contaminants in organisms. The potential for volatile compounds to be released
into the air may be an additional concern in connection with highly contaminated sites (EPA 2005a).
Release, resuspension, and production of residuals will affect risk
over different spatial scales and time frames depending on the site characteristics and nature of the dredging operation. As described by Bridges
et al. (in press), “Characterizing how dredging will influence direct risks
includes considering how the processes contributing to risk change with
time, which elements or receptors in the ecosystem are affected by these
changes, the spatial scales over which effects would be expected to occur,
and the uncertainties associated with the predicted changes and risk reduction.” As will be discussed in much greater detail in Chapter 4, resuspension and release occur in a shorter time frame during dredging
operations. Residuals will remain following dredging, however, their
distribution, longevity, and effects are poorly understood. To the extent
that release, resuspension, and production of residuals are present and
contribute risk at a site, they detract from the overall or net risk reduction resulting from the remedial activity. As such, they are an important
consideration in evaluating the effectiveness of a remediation. As noted
in the 4 Rs workshop (Bridges et al. in press) and recent EPA sediment
guidance (EPA 2005a), there is increasing recognition of the importance
of these processes and of factors that influence their control.
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rff/RFF_Press/CustomBookPages/Superfunds-Future.cfm?CFID=5819288&
CFTOKEN=75932196 [accessed Jan. 9, 2006].
Sanchez, F.F. 2001. A Multimedia Model for Assessing Chemical Fate During
Dredging of Contaminated Bed-Sediment. M.S. Thesis, Louisiana State
University.
Sanchez, F.F., L.J. Thibodeaux, K.T. Valsaraj, and D.D. Reible. 2002. Multimedia
Chemical Fate Model for Environmental Dredging. Pract. Period. Hazard.
Tox. Radioact. Waste Manage. 6(2):120-128.
Southerland, E. 2006. Overview of Superfund Sediment Remediation. Presentation at the First Meeting on Sediment Dredging at Superfund Megasites,
March 22, 2006, Washington DC.
69
Thibodeaux, L.J., and V.J. Bierman. 2003. The bioturbation-driven chemical release process. Environ. Sci. Technol. 37(13):252A-258A.
Thoms, S.R., G. Matisoff, P.L. McCall, and X. Wang. 1995. Models for alteration
of sediments by benthic organisms. Project 92-NPS-2. Alexandria, VA: Water Environment Research Foundation.
USACE (U.S. Army Corps of Engineers). 1983. Dredging and Dredged Material
Disposal. Engineer Manual 1110-2-5025. U. S. Army Corps of Engineers,
Department of the Army, Washington, DC. March 25, 1983 [online].
Available: http://www.usace.army.mil/publications/eng-manuals/em11102
5025/entire.pdf [accessed Jan. 4, 2006].
USACE /EPA (U.S. Army Corps of Engineers and U.S. Environmental Protection
Agency). 1992. Evaluating Environmental Effects of Dredged Material
Management Alternatives - A Technical Framework. EPA842-B-92-008.
U.S. Army Corps of Engineers, Department of the Army, and Office of
Water, U.S. Environmental Protection Agency, Washington, DC [online].
Available: http://www.epa.gov/owow/oceans/regulatory/dumpdredged/
framework/techframework.pdf [accessed Jan. 4, 2006].
U.S. Steel. 2004. Grand Calumet River Sediment Remediation Project Update.
U.S. Steel Gary Works Newsletter GCR Issue 6 - February 2004 [online].
Available: http://www.ussteel.com/corp/rcra/images/CITE/0402%20News
letter%20Feb%2004.pdf [accessed Aug. 3, 2006].
Washington Department of Ecology. 2005. Sediment Cleanup Status Report.
Washington State Department of Ecology Toxics Cleanup Program. Publication No. 05-09-092. June 2005 [online]. Available: http://www.ecy.wa.
gov/pubs/0509092.pdf [accessed Feb. 26, 2007].
Weston (Weston Solutions Inc.). 2002. Final Comprehensive Post-Removal Summary Report Manistique Harbor and River Site U.S. Route 2 Highway
Bridge Manistique, Michigan. Prepared for the U.S. Environmental Protection Agency, Region V. November 12, 2002.
3
Effectiveness of Environmental Dredging in
Reducing Risk: Framework for Evaluation
A wide variety of metrics can be used to evaluate the effectiveness
of environmental-dredging projects in reducing risks to human health
and the environment. The committee reviewed a number of them and
developed a framework to facilitate the evaluation of the effectiveness of
environmental-dredging projects at contaminated sediment sites. The
framework is based on the effectiveness criteria used by the U.S. Environmental Protection Agency (EPA) Superfund program to select remedies (40 CFR § 300.430[e][7][i]).
In conducting its evaluation, the committee defined dredging effectiveness as the achievement of cleanup goals defined for each site, which
take the form of remedial-action objectives, remediation goals, and
cleanup levels.1 For CERCLA remedial actions, these goals are typically
1Remedial-action objectives “are intended to provide a general description of
what remediation is expected to accomplish” (EPA 2005a, p. 2-15). For example, a
remedial-action objective might be to reduce to acceptable levels the risks to
people who ingest contaminated fish. Remediation goals are paired contaminantspecific and media-specific concentrations intended to protect human and ecologic health that incorporate site-specific information about exposure patterns and
toxicity. “At most CERCLA [Comprehensive Environmental Response, Compensation, and Liability Act] sites, [remediation goals] for human health and ecologic
70
Effectiveness of Environmental Dredging in Reducing Risk
71
documented in the record of decision (ROD). However, this definition is
appropriate for evaluating the effectiveness of the project in reducing
risk only when cleanup goals are derived from sound, site-specific risk
modeling.
An ideal evaluation of effectiveness at Superfund sites would be
based on the site conceptual model, data from a baseline assessment, and
a long-term monitoring program that permits sound statistical comparison of the spatial scale of contamination and the magnitude of risk before, during, and after dredging. At the outset of its work, the committee
hoped to obtain that kind of information for a number of large contaminated sediment sites to inform its deliberations. However, we found that
such careful and prolonged monitoring either has not been conducted,
has not been completed at large-contaminated sediment sites, or simply
was not available to the committee. (The committee noted that some sites
where remediation had not yet occurred or been completed have electronic databases and long-term monitoring plans that would facilitate
future attempts to evaluate remedy effectiveness, for example, Hudson
River and New Bedford Harbor). In some cases, it is recognized that additional information, for example, raw data and consulting reports, may
have been held by responsible parties, federal agencies, or their consultants. However, this information may not have been available in the public domain or to the committee, or the committee’s time and resource
constraints precluded a thorough compilation, analysis, and interpretation (see further discussion in Chapter 4). In some cases a review of all
site data was not necessary to determine whether cleanup goals had been
met. This chapter details the committee’s process for evaluating the effectiveness of environmental dredging within the constraints imposed by
the available data. The framework for this review is outlined in Box 3-1.
receptors are developed into final, chemical-specific, sediment cleanup levels by
weighing a number of factors, including site-specific uncertainty factors and the
criteria for remedy selection found in the NCP [National Contingency Plan]”
(EPA 2005a, p. 2-16). The ROD for each site generally should include chemicalspecific cleanup levels, indicating how these values are related to risk and how
their attainment will be measured (EPA 2005a).
72
BOX 3-1 Framework for Evaluating the Effectiveness of
Environmental Dredging Projects
1. Identify Superfund megasites and other large environmental-dredging
projects.
2. Define criteria for selecting projects for committee evaluation from the
list of environmental-dredging projects.
3. Select projects that represent a variety of site conditions.
4. Evaluate each project with respect to measures of short-term and longterm effectiveness.
5. Make recommendations for improved design, implementation, and
monitoring of future environmental-dredging projects.
SOURCES OF INFORMATION ABOUT ENVIRONMENTALDREDGING PROJECTS
The statement of task (see Appendix A) indicates that the sources of
information “would include megasites for which dredging has been
completed; megasites for which plans have been developed, partially
implemented, and operations are ongoing; and smaller sites that exhibit
lessons relevant to megasites.” The committee’s evaluation focused on
environmental-dredging projects at Superfund megasites, but, because
remediation has not been completed at many of these sites, the committee also reviewed other environmental-dredging projects on which data
relevant to the committee’s charge were available.
As described in Chapter 2, numerous environmental-dredging
projects have been conducted, and many other contaminated sediment
sites will require decision-making soon. At the first committee meeting,
EPA outlined sites from its database of tier 1 sediment sites where
dredging had been completed. From those dredging sites, EPA provided
the committee with a list of sites on which there were pre-remediation
and post-remediation monitoring data. To identify other dredging
projects for possible evaluation, the committee reviewed information
from additional government, industry, and private consulting sources
that summarize remedial activities at contaminated sediment sites (see
Box 3-2). Collectively, those dredging project compilations and reviews
73
BOX 3-2 Compilations and Reviews of Sediment Remediation Projects
EPA’s Great Lakes National Program Office summarized (EPA 1998) and
updated (EPA 2000) information on contaminated sediment sites in the Great
Lakes Basin that had been remediated with a variety of techniques. The reports
provide some details about remedies but few details on post-remedial monitoring or on whether remediation achieved expected benefits. In 1999, the Great
Lakes Water Quality Board summarized 20 sediment remediation projects in the
Great Lakes areas of concern (Zarull et al. 1999). The report stated that of the
projects implemented so far, only two (Waukegan Harbor, IL and Black River,
OH) had adequate data on long-term ecologic health. The Great Lakes binational
strategy progress reports (e.g., EPA and Environment Canada 2005) provide
annual updates summarizing sediment remediation activities at Great Lakes
sites in the United States and Canada. The Major Contaminated Sediment Sites
(MCSS) Database (GE et al. 2004) is the largest compendium of information on
such sites. It contains information on 123 major projects representing 103 sites.
EPA maintains, although not publicly, a database on the 60 tier 1 sediment sites
where remedies include dredging or excavation of at least 10,000 cy or capping
or monitored natural recovery of at least 5 acres (EPA 2005a).
Before this committee’s deliberations, several other groups reviewed data
on completed projects to examine whether remediation had been successful and
to draw conclusions about the likely effects of dredging at other sites in the future. The General Electric Company (GE 2000) evaluated sediment remediation
case studies involving 25 sites (including sites that used removal techniques
other than dredging) and attempted to determine whether data indicated the
ability to reduce risks to human health and the environment. Its report concluded that the success of the projects had not been demonstrated and that technical limitations restricted the effectiveness of dredging in reducing surface
sediment-contaminant concentrations. Cleland (2000) updated an earlier report
prepared by Scenic Hudson (2000) and outlined experience at 15 sediment remediation sites, including many evaluated by GE (2000). Cleland presented sediment and biota concentration data and concluded that post-dredging monitoring
data consistently show beneficial results, including reductions in contaminant
mass and in concentrations in sediment and fish. Many of the same case studies
were also reviewed by EPA (Hahnenberg 1999), and EPA presented the results
of its analysis to the National Research Council’s Committee on Remediation of
PCB-Contaminated Sediments (NRC 2001). EPA’s analysis indicated that dredging resulted in reduced sediment and fish-tissue concentrations. An alternative
analysis was also provided to that National Research Council committee by the
74
BOX 3-2 Continued
Fox River Group (1999, Appendix C in GE 2000), which refuted the connection
between remedial actions and fish-tissue concentration declines on the basis of a
paucity of data and flawed method in EPA’s linking of remediation to observed
changes. During its deliberations, the previous National Research Council committee (NRC 2001) reviewed the documents produced by GE (2000), Scenic Hudson (2000), EPA (Hahnenberg 1999; Pastor 1999), and the Fox River Group (1999)
to ascertain how groups reviewing the same documents could come to such disparate conclusions. It concluded: “First, in some instances, there is disagreement
about the remediation goals and the measures by which achievement of the
goals can be assessed. Second, in some cases, the available post-remediation
monitoring data are sparse and incomplete compared with pre-remediation data
and control data. Third, in some cases, it is the intention of reviewers, agencies,
and industries to support their preferences, and that might lead to more conflict.”
Since the previous committee issued its report (2001), additional groups
have sought to analyze the link between sediment remediation and reduced risk.
Baker et al. (2001) sought to address the question “Can active remediation be
implemented in such a way that it provides a net benefit to the Hudson River?”
and commented on the biologic effects of sediment remediation at five sites.
Thibodeaux and Duckworth (2001) evaluated measurements of environmentaldredging effectiveness in detail at three sites. Malcolm Pirnie (Malcolm Pirnie
and TAMS Consultants, Inc. 2004), in an appendix to the engineering performance standards developed for the Hudson River Superfund site, briefly reviewed
data on 25 remediation projects (some had not initiated remediation) and provided information on monitoring and remediation results. As part of its feasibility study for the Kalamazoo River in Michigan, BBL (2000) compiled profiles of
20 environmental-dredging projects conducted nationwide and discussed their
ability to meet project-specific objectives and their overall effectiveness. Several
of those profiles were presented in an earlier analysis (GE 2000). In addition, the
Great Lakes Dredging Team, a partnership of federal and state agencies, provided a series of dredging project case studies (Great Lakes Dredging Team
2006). The U.S. Army Corps of Engineers Center for Contaminated Sediments
also provided information on a series of environmental-dredging projects
(USACE 2006).
were extremely useful in forming the short list of dredging projects on
which pre-dredging and post-dredging monitoring data were likely to be
suitable for evaluation.
75
CRITERIA USED TO SELECT ENVIRONMENTAL-DREDGING
PROJECTS FOR EFFECTIVENESS EVALUATION
The committee used the criteria listed in Box 3-3 to select environmental-dredging projects for evaluation. Therefore, the committee preferred projects that involved only dredging but did not limit its analysis
to them. The committee did not select projects that involved dry excavation, because they are conducted under different conditions from conventional dredging. Pilot studies, although limited in scope and purpose,
were also considered because they are often heavily monitored with substantial information regarding the effect of dredging under specified
conditions.
The committee preferred projects with removal of at least 10,000 cy
of sediment because of their similarity to megasites with respect to the
spatial scale of ecologic and human health exposures. Some smaller pilot
studies designed to inform decisions on larger sites were also included.
This size threshold matches the threshold used by EPA to define tier 1
sediment sites (EPA 2005a).
Any evaluation of dredging effectiveness depends on sufficient
high-quality pre-dredging and post-dredging monitoring data. Therefore, another criterion used by the committee to select dredging projects
was whether they had pre-dredging and post-dredging monitoring data.
Environmental-dredging projects occur in a wide variety of aquatic
environments, such as rivers, estuaries, bays, lakes, ponds, canals, and
wetlands; and the effectiveness of dredging can be influenced by many
BOX 3-3 Criteria Used to Select Environmental-Dredging
Projects for Evaluation
1. The remedy consists of dredging only or a combined remedy that includes dredging.
2. The remedy preferably includes removal of at least 10,000 yds3 of sediment.
3. The project has some amount of pre-dredging and post-dredging data.
4. The projects collectively represent a wide variety of project types (for
example, environmental settings, chemicals of concern, and dredging design and
implementation).
76
site-specific factors. Therefore, the committee endeavored to conduct a
broad review of projects that represented a variety of site-specific conditions, including the type of water body, the form of chemical contamination, and the type of dredging technology.
SELECTION OF ENVIRONMENTAL-DREDGING PROJECTS
The committee chose the 26 dredging projects listed in Table 3-1 for
detailed evaluation (see Figure 3-1 for the location of the projects).
Selected projects involved remedies with dredging only or dredging
combined with backfilling or capping. Chemicals of concern included
polychlorinated biphenyls (PCBs), polycyclic aromatic hydrocarbons
(PAHs), pesticides, and metals detected in sediments in a variety of
aquatic environments across the United States and one in Sweden. PCBs
were the primary chemicals of concern in 19 of the 26 projects. Table 3-1
contains general information on the site and the remedy; further detail
on several of the sites is provided in Chapter 4 and the associated
references. In addition, summaries of many of these sites and their
remediation have been compiled elsewhere (e.g., GE 2004; Malcolm
Pirnie, Inc. and TAMS Consultants, Inc. 2004).
The selected projects include all 12 sites (although not all the operable units at each site) on which there were pre-dredging and postdredging data as reported by EPA at the first committee meeting (Ells
2006). The committee also chose three demonstration projects (two in the
Lower Fox River, WI, and one in the Grasse River, NY) and one statelead site (Cumberland Bay, NY) recommended for review by the Sediment Management Work Group during the committee’s second meeting
(Nadeau 2006).
Scenic Hudson (Cleland 2000), an advocacy group that promotes
cleanup of the Hudson River in New York, also included those demonstration projects and Cumberland Bay in its review of dredging effectiveness. In addition, Scenic Hudson identified Lake Jarnsjon in Sweden
and the Black River, OH, as examples of successful dredging operations
(as did Green and Savitz [unpublished] and Zarull et al. 1999), and the
committee selected these two sites for evaluation. During the committee’s second meeting, remediation consultants presented results of two
recently completed large-scale dredging operations: in the Lower Fox
Estuary
Hg
PAHs
PCBs
PCBs, As,
PAHs
As, Cu
PCBs
Lavaca Bay, TX
Black River, OH
Outboard
Marine Corp.,
Waukegan
Harbor, IL
Commencement
Bay–Head of
Hylebos,
Tacoma, WA
Commencement
Bay–Sitcum,
Tacoma, WA
Duwamish
Diagonal,
Seattle, WA
Tidally influenced
river
Estuary
Estuary
Harbor
River
Bayou
Water- Body Type
Bayou Bonfouca, PAHs
LA
Project
Primary
Chemicals of
Concern
Dredging followed by
capping
Dredging
Dredging
Dredging
Dredging
Dredging
Dredging followed by
backfilling
Remedy Type
Full-scale
Full-scale
Full-scale
Full-scale
Full-scale
Pilot
Full-scale
Scale of
Effort
TABLE 3-1 Summary of 26 Environmental-Dredging Projects Selected for Evaluationa
2003-2004
1993-1994
2003-2006b
1991-1992
1989-1990
1998
1994-1995
Dates of
Dredging
66,000
428,000d
419,000c
38,000
45,000-60,000
80,000
170,000
Volume of Dredged
Sediment (cy)
77
As, Pb, Zn, Cu,
PAHs, PCBs,
tributyltin, Hg
PCBs
Zn, Pb, Cd
Harbor Island–
Todd, Seattle,
WA
Cumberland
Bay, NY
Dupont,
Christina River,
DE
River
River
Lower Fox River PCBs
(OU-1), WI
(Deposit N), WI
River
Inland lake
Estuary
PCBs, PAHs,
Estuary
Hg, Pb, As, Cu,
Zn, tributyltin
Harbor Island–
Lockheed,
Seattle, WA
Estuary
PCBs
Water- Body Type
Puget Sound
Naval Shipyard,
Bremerton, WA
Project
Primary
Chemicals of
Concern
TABLE 3-1 Continued
Full-scale
Scale of
Effort
Dredging
Dredging
Dredging and backfilling
Dredging
Dredging, capping, and
habitat restoration in
selected areas
Pilot
Full-scale
Full-scale
Full-scale
Full-scale
Dredging in open-water
Full-scale
area, dredging and
capping in nearshore area
Dredging
Remedy Type
1998-1999
2004-present
1999
1999-2000
2004-2005
2003-2004f
2000-2004
Dates of
Dredging
8,200
Incomplete
11,000
195,000
220,000
70,000
225,000e
Volume of Dredged
Sediment (cy)
78
Fjord
Bay
Ketchikan Pulp 4-methyl
Company, Ward phenol;
Cove, AK
ammonia
PAHs, PCBs
PCBs
PCBs
PCBs
PCBs
Newport Naval
Complex–
McCallister
Landfill, RI
GM Central
Foundry, St.
Lawrence River,
NY
Grasse River,
NY (non-timecritical removal
action)
Grasse River,
NY remedial
options pilot
study (ROPS)
Lake Jarnsjon,
Sweden
Lake
River
River
River
River
(SMU 56/57), WI
Dredging
Dredging
Dredging (backfilling in
one area)
(thin-layer)
Full-scale
Pilot
Pilot
Full-scale
Full-scale
Full-scale
Pilot
1993-1994
2005
1995
1995
2001
2000-2001
1999-2000
196,000g
30,000
3,000
14,000
34,000
8,700
82,000
79
Dredging
Dredging
Dredging
(backfilling in one area)
Dredging
Remedy Type
Full-scale
Full-scale
(hot spot
removal)
Full-scale
Full-scale
Full-scale
Scale of
Effort
1996-1997
1994-1995
1993-1995
2001
1995-2000
Dates of
Dredging
110,000
14,000
71,000
86,000
190,000h
Volume of Dredged
Sediment (cy)
aIn the review and collection of data on dredging sites it was determined that various sources will often provide inconsistent information. For
example, the committee found several different estimates for the volume of sediment removed from the Black River: 49,700 cy (Zarull et al. 1999;
Cleland et al. 2000); 45,000 cy (Malcolm Pirnie and TAMS Consultants, Inc. 2004); 50,000 cy (EPA 2000); 60,000 cy (GE 2000; GE 2004; Fox R. Group
1999).
bFrom Dalton, Olmsted & Fuglevand, Inc. 2006.
DDT, dieldrin
Estuarine channel
River
United
Heckathorn,
Richmond, CA
Cd
Marathon
Battery, Hudson
River, Cold
Spring, NY
River
Estuary
PCBs, PAHs,
total
dibenzofuran
Reynolds
Metals, St.
Lawrence River,
NY
Harbor
Water- Body Type
New Bedford
PCBs
Harbor, MA (hot
spot)
PCBs
Manistique
Harbor, MI
Project
Primary
Chemicals of
Concern
TABLE 3-1 Continued
80
15,000 cy removed with upland-based equipment and 404,000 removed with marine-based equipment.
From EPA summary (EPA 2006 [Commencement Bay–Sitcum Waterway, April 26, 2006]): “Approximately 396,000 cy of sediment were dredged
from Area A using hydraulic and clamshell dredges and approximately 32,300 cy were removed from Area B using a small hydraulic dredge.”
eFrom EPA summary (EPA 2006 [Puget Sound Naval Shipyard, May 15, 2006]).
fFrom EPA summary (EPA 2006 [Harbor Island Lockheed Shipyard Sediment Operable Unit, May 11, 2006]): “Remediation dates: These dates
refer to dredging activities, not other remedial activities such as capping. November 22, 2003 to March 10, 2004 and from October 22, 2004 to
November 22, 2004.”
gBremle et al. 1998.
hEPA 2006 (Manistique River and Harbor Site, May 10, 2006).
Sources: Data are also unpublished data from EPA, NAS Completed Dredging Summary, March 22, 2006; Malcolm Pirnie, Inc., and TAMS
Consultants, Inc. 2004.
d
c
81
82
FIGURE 3-1 Locations of environmental-dredging projects selected for evaluation. Several locations comprise more than one site, including Commencement
Bay, Grasse River, Lower Fox River, and Harbor Island.
River Operable Unit 1 and the Grasse River Remedial Options Pilot
Study (Connolly et al. 2006; Fox et al. 2006). Those two projects were selected for evaluation because extensive monitoring data on them were
made available to the committee. The committee selected a pilot dredging study in Lavaca Bay, TX, and a hot spot removal in New Bedford
Harbor, MA, because they were the subject of extensive monitoring efforts. The Reynolds Metals Superfund site in the St. Lawrence River was
selected because it has undergone monitoring for several years after initial dredging (EPA 2005b).
Of the 26 dredging projects, five have been identified by EPA as
contaminated sediment megasites, that is, sites where the dredging portion of the remedy will cost at least $50 million (see “Sediment Contamination at Superfund Sites” in Chapter 2).
83
APPROACH TO EVALUATING DREDGING PROJECTS
The committee evaluated the effectiveness of the 26 dredging projects, where data permitted such an evaluation, by determining whether
cleanup levels and remedial action objectives were achieved. The committee used lessons learned from the projects to identify means for improving future decisions about remediation at Superfund megasites. The
lessons learned involved:
▪
▪
▪
ness.
Factors that contributed to the success of dredging operations.
Factors that adversely affected dredging operations.
Methods for improving the monitoring of dredging effective-
The committee did not attempt to substitute its own judgment
about what remedial action objectives and cleanup levels should be, including the site-specific risk modeling on which they were based, but
simply tried to determine whether the stated goals were achieved and
why or why not. The committee chose to review sites where dredging
was the only remedy (or at least the main remedy) to evaluate the effectiveness of only dredging. However, the committee acknowledges that
the effects of dredging may be difficult or impossible to distinguish from
other ongoing processes including additional source control and natural
recovery.
The committee recognizes the importance of quantifying the net
risk reduction associated with dredging, which involves consideration of
possible increases in risk to the spectrum of human and environmental
receptors that might be caused by dredging and the treatment, storage,
transport, and disposal of dredged sediment as well as risk reduction
from removal of contaminated sediments from aquatic systems. The
need for considering net risk reduction and incorporating the analysis
into decision-making at Superfund sites has been noted in previous National Research Council reports (NRC 2001; NRC 2005), recognized by
EPA (EPA 2005a), and examined in the scientific literature (Wenning et
al. 2006). However, such analysis was not possible in this case. It was
difficult (in some cases impossible) to obtain requisite data on the dredging component of each project, much less some of the other measures
relevant to a net risk reduction analysis (for example, number of acci-
84
dents during transportation of dredged material) that are often not compiled. These limitations combined with time and resource constraints
prevented the committee from conducting net risk reduction evaluations
for each dredging project. As a result, the committee chose to focus its
time and resources on the dredging component of projects rather than
treatment, storage, transport, and disposal.
Methods Used to Evaluate Dredging Projects
Each project review began with the identification of the remedial
action objectives and cleanup levels for the project. That was followed by
review of project data to judge whether the remedial action objectives
and cleanup levels had been met. Those data could include chemical
concentrations in sediment, surface water, the biota, and other environmental media; indicators of exposure (such as bioaccumulation testing
and human biomonitoring data); and measures of risk (such as toxicity
testing). The measures have inherent natural heterogeneity and are subject to uncertainty. When, despite the variability and uncertainty, sufficient data were available to compare pre-dredging and post-dredging
conditions, the committee attempted to answer the effectiveness questions listed in Box 3-4. Ideally, the evaluation should be performed in the
context of a comparison of all available remedial alternatives (EPA
2005a; Bridges et al. 2006), but the committee reviewed only dredging
projects in accordance with the statement of task.
Cleanup levels and remedial action objectives are not always riskbased, and some projects have no cleanup levels or other quantitative
means to judge effectiveness at all. Therefore, the committee looked beyond remedial action objectives and cleanup levels to identify dredging
successes and failures in reducing exposure or risk, as well as the site,
the remedy, or the contaminant conditions that led to the successes and
failures. In its evaluation of project data, the committee distinguished
between changes in environmental-media concentration, exposure, and
risk.
The committee evaluated whether baseline assessment data were
available to define conditions before dredging for comparison with
monitoring data collected during and after dredging. Optimally, those
85
BOX 3-4 Measures of Sediment-Remedy Effectiveness
in the Superfund Program
Dredging effectiveness is the achievement of cleanup goals (remedial action
objectives and cleanup levels that are derived from appropriate site-specific risk
modeling) in the predicted time frame with a reasonable degree of confidence
that the achievement will be maintained.
Interim Measures
1. Short-term remedy performance. For example, have sediment cleanup
levels been achieved after dredging?
2. Long-term remedy performance. For example, have sediment cleanup
levels been maintained for at least 5 years, and thereafter as appropriate?
3. Short-term risk reduction. For example, have remedial action objectives
been achieved? Do data demonstrate or at least suggest a reduction in fish tissue
concentrations, a decrease in benthic toxicity, or an increase in species diversity
or other community indexes after 5 years?
Key Measure
4. Long-term risk reduction. For example, have remedial action objectives
been maintained for at least 5 years, and thereafter as appropriate?
5. Has the predicted magnitude and timing of risk reduction been
achieved or are they likely to be achieved?
assessment data would suffice to quantify exposures and risks and allow
comparison with during-dredging and post-dredging monitoring data.
Human and ecologic exposure and risk might increase during dredging,
and these increases should also be weighed against exposure and risk
reductions following dredging. Monitoring data and information collected during dredging were reviewed to identify changes in human and
ecologic exposure and risk that occurred during this period. Dredging of
contaminated sediment disrupts the bottom substrate, thereby destroying the existing benthic community, and can increase exposure of humans and the biota, depending on the degree of resuspension, residual
generation, and release of sediment-bound, dissolved, or airborne con-
86
tamination. The committee also reviewed monitoring data and information collected after dredging to identify changes in human and ecologic
exposure and risk that resulted from dredging and to evaluate whether
the changes should be expected to be maintained despite extreme
weather conditions and human activities.
The committee used lessons learned from individual dredging project evaluations to inform its deliberations about how to improve future
remediation decisions at Superfund megasites. Specifically, the committee sought to define site conditions and project design implementation
factors that affect dredging success and to use this information in recommending improved management and monitoring to facilitate scientifically based and timely decision-making.
REFERENCES
Baker, J.C., W.F. Bohlen, R. Bopp, B. Brownawell, T.K. Collier, K.J. Farley, W.R.
Geyer, and R. Nairn. 2001. PCBs in the Upper Hudson River: The Science
behind the Dredging Controversy. A White Paper Prepared for the Hudson River Foundation, New York, NY. October 25, 2001 [online]. Available:
http://www.seagrant.sunysb.edu/HEP/archive/hrfpcb102901.pdf [accessed
Aug. 16, 2006].
BBL (Blasland, Bouck & Lee, Inc). 2000. Allied Paper, Inc./Portage
Creek/Kalamazoo River Superfund Site RI/FS Feasibility Study Report—
Phase I Allied Paper, Inc., Appendix D. Site Profiles of Sediment Dredging
Projects. Draft for State and Federal Review. October 2000 [online]. Available: http://www.deq.state.mi.us/documents/deq-erd-kzoo-FS-apend-d.pdf
[accessed Jan. 8, 2006].
Bremle, G., L. Okla, and P. Larsson. 1998. PCB in water and sediment of a lake
after remediation of contaminated sediment. Ambio 27(5):398-403.
sediments. Integr. Environ. Assess. Manage. 2(1): 51-58.
Cleland, J. 2000. Results of Contaminated Sediment Cleanups Relevant to the
Hudson River: An Update to Scenic Hudson's Report “Advances in Dredging Contaminated Sediment.” Prepared for Scenic Hudson, Poughkeepsie,
NY. October 2000 [online]. Available: http://www.scenichudson.org/pcbs/
pcb_dredge.pdf [accessed Aug. 16, 2006].
Connolly, J., V. Chang, and L. McShea. 2006. Grasse River Remedial Options
Pilot Study (ROPS) Findings from Dredging Activities. Presentation at the
87
Second Meeting on Sediment Dredging at Superfund Megasites, June 7,
2006, Irvine, CA.
Ells, S. 2006. Summary of Available Information on Superfund Sediment Sites
and Dredging Sites. Presentation at the First Meeting on Sediment Dredging at Superfund Megasites, March 21-22, 2006, Washington, DC.
EPA (U.S. Environmental Protection Agency). 1998. Realizing Remediation: A
Summary of Contaminated Sediment Remediation Activities in the Great
Lakes Basin. U.S. Environmental Protection Agency, Great Lakes National
Program Office, Chicago, IL. March 1998 [online]. Available: http://www.
epa.gov/glnpo/sediment/realizing/realcover.html [accessed Aug. 16, 2006].
EPA (U.S. Environmental Protection Agency). 2000. Realizing Remediation II: A
Summary of Contaminated Sediment Remediation Activities at Great
Lakes Areas of Concern. U.S. Environmental Protection Agency, Great
Lakes National Program Office, Chicago, IL. July 2000 [online]. Available:
http://www.epa.gov/glnpo/sediment/realizing2/index.html [accessed Aug.
16, 2006].
http://www.epa.gov/superfund/resources/sediment/guidance.htm
EPA (U.S. Environmental Protection Agency). 2005b. Five-Year Review Report
Reynolds Metals Company Site St. Lawrence County Town of Massena,
New York. U.S. Environmental Protection Agency, Region 2, New York.
April 2006 [online]. Available: http://www.epa.gov/superfund/sites/five
year/f06-02018.pdf [accessed Jan. 8, 2007].
EPA (U.S. Environmental Protection Agency). 2006. Case Study Data and Information. U.S. Environmental Protection Agency.
EPA and Environment Canada (U.S. Environmental Protection Agency and Environment Canada). 2005. Great Lakes Binational Toxics Strategy: 2005
Annual Progress Report. U.S. Environmental Protection Agency, Chicago,
IL, and Environment Canada, Environmental Protection Branch, Downsview, ON [online]. Available: http://binational.net/bns/2005/2005-GLBTSEnglish-web.pdf [accessed Sept. 12, 2006].
88
Fox River Group. 1999. Effectiveness of Sediment Removal. September 27, 1999
(Appendix C in GE 2000)
Fox, R., M. Jury, and J. Kern. 2006. Overview of Lower Fox River OU1 Dredge
Residuals. Presentation at the Second Meeting on Sediment Dredging at
Superfund Megasites, June 7, 2006, Irvine, CA.
GE (General Electric Company). 2000. Environmental Dredging: An Evaluation
of Its Effectiveness in Controlling Risks. General Electric Company,
Blasland, Bouck & Lee, Inc., and Applied Environmental Management,
Inc., Albany, New York. August 2000 [online]. Available: http://www.ge.
com/files/usa/en/commitment/ehs/hudson/DREDGE.pdf [accessed Jan. 8,
2006].
GE (General Electric Company). 2004. Major Contaminated Sediments Sites
(MCSS) Database, Release 5.0. General Electric Company, Applied
Environmental Management, Inc., and Blasland, Bouck & Lee, Inc.
September 17, 2004 [online]. Available: http://www.ge.com/en/citizenship/
ehs/remedial/hudson/mcss/database.htm [accessed Jan. 8, 2006].
Great Lakes Dredging Team. 2006. Case Study Series [online]. Available:
http://www.glc.org/dredging/ [accessed Sept. 12, 2006].
Hahnenberg, J. 1999. Long-term benefits of environmental dredging outweigh
short-term impacts. Engineering News Record (March 22/29):E15-E16.
Malcolm Pirnie, Inc, and TAMS Consultants, Inc. 2004. Final Engineering Performance Standards Hudson River PCBs Superfund Site, Appendix: Case
Studies of Environmental Dredging Projects. Prepared for U.S. Army
Corps of Engineers, Kansas City District, by Malcolm Pirnie, Inc., White
Plains, NY, and TAMS Consultants, Inc., Bloomfield, New Jersey. April
2004 [online]. Available: http://www.epa.gov/hudson/eng_perf/FP5001.pdf
[accessed Aug. 16, 2006].
Nadeau, S. 2006. SMWG Review and Analysis of Selected Sediment Dredging
Projects. Presentation at the Second Meeting on Sediment Dredging at
NRC (National Research Council). 2005. Superfund and Mining Megasites: Lessons from the Coeur d’Alene River Basin. Washington, DC: The National
Academies.
Pastor, S. 1999. Dredging: Long-term benefits outweigh short-term impacts. Fox
River Current 2(4):7-8 [online]. Available: http://www.epa.gov/Region5/
sites/foxriver/current/foxcurrent199909.pdf [accessed Sept. 12, 2006].
Scenic Hudson. 2000. Accomplishments at Contaminated Sediment Cleanup Sites
Relevant to the Hudson River. Scenic Hudson, Poughkeepsie, NY. September 2000.
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Thibodeaux, L.J., and K.T. Duckworth. 2001. The effectiveness of environmental
dredging: A study of three sites. Remediation J. 11(3):5-34.
USACE (U.S. Army Corps of Engineers). 2006. Case Studies and Projects [online].
Available: http://el.erdc.usace.army.mil/dots/ccs/case.html [accessed Sept.
12, 2006].
Zarull, M.A., J.H. Hartig, and L. Maynard. 1999. Ecological Benefits of Contaminated Sediment Remediation in the Great Lakes Basin. Windsor, Ontario:
International Joint Commission [online]. Available: http://www.ijc.org/
php/publications/html/ecolsed/ [accessed Sept. 12, 2006].
4
Evaluation of Dredging Effectiveness:
What Has Experience Taught Us?
INTRODUCTION
Over the last 20 years, various contaminated sediment sites have
been remediated in whole or in part through dredging. The committee
examined 26 dredging projects to evaluate whether dredging was able to
meet short-term and long-term goals. Short-term goals are defined as
cleanup levels that can be measured during or immediately postdredging to verify the effective implementation of the remediation. The
ability to maintain cleanup levels in the long term is ideally linked to the
achievement of long-term risk-based goals or remedial action objectives.
Appendix C presents the various sites’ cleanup levels and remedial action objectives and describes whether they were achieved at individual
sites. Taken as a whole, the projects indicate what can and cannot be
achieved with dredging and the conditions that favor or discourage the
use of dredging.
Evidence that dredging projects led to the achievement of longterm remedial action objectives and did so within expected or projected
time frames is generally lacking. It was often not possible to evaluate
long-term remedy performance relative to remedial action objectives because of insufficient post-remediation data, quality, or availability or because of lack of an equivalent pre-remediation dataset. Post-remediation
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Evaluation of Dredging Effectiveness
91
conditions are always influenced by long-term natural attenuation processes and ongoing sources of contaminations if they exist, so long-term
monitoring—over decades—may be needed to establish effectiveness;
few of the sites reviewed by the committee have reached that level of
maturity. Not counting the 5 pilot studies or hot-spot removal actions,
about one half of the sites apparently did not achieve remedial action
objectives or had inadequate monitoring to judge performance relative to
remedial action objectives. Insufficient time has elapsed to judge
achievement of remedial action objectives in approximately one quarter
of the sites. The remaining sites apparently met remedial action objectives although the extent to which those remedial action objectives
achieve long-term risk reduction may not be known.
There were often sufficient data to evaluate performance relative to
cleanup levels or short-term implementation goals, but the relationship
of these measures to long-term risk reduction was often not clear. An
examination of Appendix C shows that many sites achieved cleanup levels; however, many were operational goals (mass removal or dredging to
elevation) rather than contaminant-specific goals.
Natural processes are always modifying conditions at a site; their
influence can be difficult or impossible to separate from the remedial
action, particularly when control or reference sites are not monitored
before and after remediation. Conditions also are often influenced by the
implementation of combined remedies, such as dredging and capping,
which complicate the assessment of the performance of dredging alone.
Thus, the committee was unable to evaluate the effectiveness of dredging
alone at most sites.
Experiences at the sites can nevertheless inform remedial project
managers as to what may be achievable with dredging and what site and
operational factors may limit dredging effectiveness or contribute to its
success. Experience is especially useful in identifying factors that contribute to success or failure of dredging to meet short-term cleanup levels
because monitoring has been conducted at most sites to judge performance relative to these standards. The ability to meet chemical-specific
cleanup levels, however, does not in itself mean the ability to meet longterm risk-reduction targets or indicate the time frame over which any
such targets might be met. This chapter discusses the lessons learned
from sites where dredging was conducted and uses specific examples to
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illustrate them. It also provides recommendations regarding implementation of successful remediation with dredging.
DATA AVAILABILITY AND ACCESSIBILITY
The potential utility of a review of remedial effectiveness is governed by the availability of pre-remediation and post-remediation monitoring data. That issue has three components: whether data were acquired at a site, whether the data are available for review, and whether
the data are sufficient to support conclusions about effectiveness. The
goal of acquiring data for this type of analysis appears relatively simple:
collect and evaluate pre-remediation and post-remediation monitoring
data on concentrations and effects from Superfund sites. However, obtaining this information is surprisingly difficult.1,2
For example, the post-dredging sediment concentrations at the Waukegan
Harbor, IL, Superfund site were of interest. In response to the committee’s request for these data, the U.S. Environmental Protection Agency (EPA) stated only
“Post-Cleanup: 1/1/2005, Ave. 2.5 ppm (Source: RPM),” with no supporting data
(EPA 2006a [OMC Waukegan Harbor Site, May 15, 2006]). To pursue sediment
data further, the committee reviewed the site’s 5-year reviews (EPA 1997a, 2002).
The 1997 5-year review states that “confirmation sampling was taken at the base
of dredge to verify that contaminants levels required for this cleanup were met.”
However, the sampling did not include chemical analyses of contaminant levels
in the sediment (Canonie Environmental 1996). The 2002 5-year review does not
indicate that any post-remediation monitoring of the harbor sediments had been
conducted. However, another literature search indicated that EPA contaminants
studied in harbor sediments collected in 1996, a few years after dredging (EPA
1999) and again in 2003 (ILDPH/ATSDR 2004).
2For example, the post-dredging sediment concentrations at the Bayou Bonfouca, LA, Superfund site were of interest. In EPA’s summary to the committee
(EPA 2006a [Bayou Bonfouca Superfund site, May 12, 2006]), they indicate that
COC concentration data in sediment and biota are unavailable. The “monitoring”
section in EPA’s summary refers only to a study by the Hazardous Substance
Research Center S&SW which contains an analysis of remedy performance but
not the raw monitoring data. However, the conclusion contains an excerpted
paragraph from a 2003 report on the site (EPA 2003a) that indicates that postdredging sediment samples were collected (a portion of the 2003 report [without
sample locations or relation to the dredging site] was provided). The locations
were later received from EPA. To pursue obtaining sediment data further, the
1
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The amount, frequency, and type of data collected at dredging sites
are highly variable. Some of the earlier sites had very little post-dredging
monitoring. For example, Bayou Bonfouca, LA, and Outboard Marine
Corporation, IL, did not sample sediment concentrations immediately
after dredging (see footnotes 1 and 2). Marathon Battery, NY, is an exception in that sediment and biota concentrations were collected and
bioaccumulation was tested, but obtaining monitoring data proved difficult, requiring several iterations, and ultimately the committee could not
access the full range of reports. At some more recent sites, dredging is
supported and guided by chemical confirmation sampling, and the resulting data are accessible. For example, the U.S. Environmental Protection Agency (EPA) provided the committee concentration and location
data on the recently completed dredging in the Grasse River, NY, and
Hylebos Waterway in Commencement Bay, WA, including date, location
coordinates, and chemical concentration information. Information on the
operations and sampling and the monitoring results at pilot projects
(such as conducted at the Grasse River, NY; Fox River, WI; and Lavaca
Bay, TX, sites) are often well documented, as would be expected from
studies specifically intended to document remedial effectiveness on a
smaller scale.
Data and reports from remediation sites are often held by various
entities (including EPA, consultants, states, and responsible parties), and
site’s 5-year reviews (EPA 1996a, 2006b; CH2M Hill 2001) were examined. The
2001 5-year review (the first to address the dredging activity) states that “no
monitoring of the water level or quality conditions in the bayou are currently
conducted—and no water quality data has been collected in the bayou adjacent
to the site since the end of the source removal remedial action in 1995.” However,
it also states that “the swimming and sediment contact advisory remains in effect
based on the sediment samples collected [by the State of Louisiana] in 1997.” The
5-year review (CH2M Hill 2001) does not provide these sampling data or indicate
locations or average concentrations of the samples. Later efforts to obtain the
data through EPA were not successful. The 2006 5-year review indicates that
sediment samples were taken after Hurricane Katrina in 2005. The post-hurricane
sampling report was not provided to the committee, but the data (without sample locations) were provided in the 2006 5-year review. The post-hurricane report
(CH2MHILL 2006), with sample locations, was later requested and acquired
from EPA.
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this complicated the compilation of information. Committee requests for
data on sites were sent to Superfund Headquarters, which did not have
them and forwarded the requests to project managers in the EPA region
who were responsible for a particular site (although these managers may
or may not have been in that position during the remediation). The data
might not be held by the EPA region, but instead may reside with the
contractors that performed the work or the responsible parties that
funded the work. Some data at sites where remedial actions had been
completed are archived or not readily retrievable. Thus, even when information was available to EPA it might have been inaccessible. In some
cases, reports containing monitoring data and interpretations were held
by the responsible parties but EPA wished not to have them released
because sensitive negotiations were under way.
When reports and data were available, they may have been reproduced only on paper although they were originally produced electronically. Such conditions severely limit distribution and faithful replication
of information, because many site documents rely on large-scale maps in
color. The ability to access reports and data via public Web sites was
generally extremely limited, but there were exceptions. The mid-Atlantic
EPA Region 3 has each site’s administrative records on line, and this
permits the public and researchers to access site files electronically (although typically these files are in a scanned, nonnative format).3 Public
information is available on all Superfund sites via the CERCLIS database,
which frequently contains a site’s record of decision and 5-year reviews,
if available. It was presumed that a site’s 5-year review would contain
explicit statements of the sampling that had been conducted and provide,
at a minimum, concentration and location data on sampling, but the
committee was surprised to see that that was not necessarily the case.1,2
Comments regarding the ready public accessibility of electronic
data may seem trivial. However, pre-remediation and post-remediation
information is the end result of massive planning, implementation, and
The administrative records contain much information that is ancillary to understanding site conditions (for example, e-mails, records of phone discussions,
submitted comments, and written communications among states, agencies, and
responsible parties). Those materials are important for maintaining a transparent
decision-making process, but the primary data reports and summaries are most
important for reviewing remedial effectiveness.
3
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data collection efforts that typically have involved large expenditures of
public resources (whether the expenditures result from remediation itself
or from the establishment of agreements with responsible parties who
conduct it). Provision of pre-remediation and post-remediation data on
chemicals of concern in an accessible, intuitive manner that defines collection efforts and results is a prerequisite to reviewing and understanding the results of remedial projects.
An issue related to the availability (or lack) of complete sampling
data is the need to rely on data summaries and various site reports to
evaluate pre- and post-remediation results. At times the committee relied
on reports and summaries that did not convey the necessary raw data to
confirm summary statistics. That is because the committee did not have
access to the primary sources or the resources to complete an ad-hoc reassembly and evaluation of all the information. A note of caution relevant to this and other studies that summarize site information is that
data on concentrations and effects should be collected consistently over
time (for example, from the same locations, media, depth interval, and
developed with similar techniques and protocols) to be most useful.
Summary statistics may not be derived from similar datasets and reports
may have incomplete annotation on sample location (for example, the
relation of the samples to the dredging footprint), sampling protocols,
and chemical analyses. Over time, analytical methods, contractors, sampling locations, and sampling methods can change. These changes complicate pre- and post-remediation comparisons. When possible, the
committee provides information on these issues.
DREDGING EFFECTIVENESS
Dredging to Remove Contaminant Mass
The direct effect of dredging is the removal of sediment and its associated contaminant mass. Experience at a variety of sites has shown
that dredging is effective at removing contaminant mass. Where sediments are subject to scour by storm or other high-flow events, buried
contaminated sediment may be the source of future exposure and risk. In
such cases, mass removal may result in risk reduction because the future
96
exposure and transport of sediment have been thwarted (see Chapter 2
for additional discussion).
For example, a demonstration dredging project was conducted to
remove a deposit (Deposit N) contaminated with polychlorinated biphenyls (PCBs) in a high-velocity reach of the Lower Fox River, WI, in
1998 and 1999. PCB contamination in sediments of the river is the result
of historical wastewater discharge from the manufacture and recycling of
carbonless copy paper incorporating Aroclor 1242. The objective of the
Deposit N demonstration was to remove contaminated sediment and
leave no more than 3-6 in. of residual material in place while minimizing
resuspension and offsite loss of sediment. Dredging to target elevations
in 1998 and 1999 resulted in the removal of 112 kg of PCBs, or 78% of the
pre-dredging inventory (Foth and Van Dyke 2000). Mass removal may
have been an appropriate cleanup objective if there was the potential for
future mobilization and transport of the PCBs.
Simple mass removal, however, may not reduce risk. For example,
the non-time-critical removal action conducted in 1995 in the Grasse
River in Massena, NY, had the objective of removing much of the PCB
mass that was in the vicinity of an outfall. PCBs had been in use at the
adjacent Alcoa facility and were introduced into the river through the
outfall and from other sources. It was estimated that this localized removal of about 2,500 cy removed 27% of the PCB mass from the entire
study area, consisting of several miles of river (BBL 1995). Despite removal of as much as 98% of the targeted contaminant mass (Thibodeaux
and Duckworth 1999), no measurable reduction in water-column or fish
concentrations of PCBs was noted. Site characterization and assessment
efforts have led to the conclusion that water-column PCB concentrations
are related, at least during low-flow periods, to surficial sediment concentrations of PCBs throughout the river and that removal of buried
mass does not have a major influence on water-column concentrations
(Ortiz et al. 2004). The removal may still have been warranted to avoid
potential scouring during high-flow conditions, but risk reduction was
not achieved during base flow conditions.
Dredging to Reduce Risk
A more complete assessment of dredging effectiveness would include evaluation of long-term risk reduction in addition to mass removal
97
or performance relative to cleanup levels. Although few sites have sufficiently complete datasets, dredging has apparently resulted in risk reduction in some cases, including at some sites with long-term datasets.
Lake Jarnsjon, Sweden
The Lake Jarnsjon site was remediated in 1993 and 1994. The predredging surface sediment (0-40 cm) PCB concentrations had a geometric
mean of 5.0 mg/kg (n=12; range 0.4 to 30.7 mg/kg) in 1990. Following
remediation, the surface sediment (0-20 cm) concentrations in 1994 were
significantly reduced and had a geometric mean of 0.060 mg/kg (n=54;
range 0.01 to 0.85 mg/kg) (Bremle et al 1998a). Out of 54 defined subareas,
one exceeded 0.5 mg/kg dry weight (set as the highest acceptable level to
be left in the sediment); 20% of the sediment areas had PCB levels higher
than 0.2 mg/kg dry weight, the remediation objective was set at 25%
(Bremle et al 1998b).
Fish-tissue PCB concentrations declined after remediation although
post-dredging monitoring did not take place until 2 years after dredging.
Concentrations did not, however, decrease to those upstream of the contaminated area. In their report, Bremle and Larsson (1998) compare fish
concentrations in the remediated lake to fish in upstream areas and conclude that “fish from all the locations in 1996 had lower PCB concentration than in 1991 [dredging occurred in the summers of 1993 and 1994].
The most pronounced decrease was observed in the remediated lake,
where levels in fish were halved. The main reason for the reduced levels
was the remediation.” However, the authors also state that “the reason
for the decline of PCB in fish could be decreased atmospheric deposition
and thus lowered loadings of PCB to the freshwater.”
The comparisons in the study benefited from the use of a reference
site that indicated background declines in fish-tissue concentrations;
these declines have been seen elsewhere as well (Stow et al. 1995). In that
regard, Bremle and Larsson (1998) state that
…the results show that if a remedial action is to be evaluated and
the process is extended over several years, changes in background
contamination must be taken into account. After a remedial action,
the results need to be followed over several years to show if it has
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been successful, which has not yet been the case in the present
study. It also stresses the importance of using reference sites, to
compare the results from the remedial area. A decrease in overall
background contamination could otherwise well be interpreted as a
result of the remedial action only.
Black River, Ohio
At the Black River, in Lorain, OH, sediment was contaminated with
polycyclic aromatic hydrocarbons (PAHs) as a result of effluent from a
steel-plant coking facility. In 1983, the coking facility closed. Dredging
occurred from late 1989 to early 1990 below the Kobe Steel outfalls at
river miles 2.83-3.55 (EPA 2007a). In the early 1980s, PAH compounds
were detected in sediments at high concentrations, and the brown bullhead population had high rates of liver cancer and pre-cancerous lesions.
Since closure of the facility and dredging, PAH concentrations in surface
sediments, fish PAH residues, and neoplasm frequencies in fish have
declined (Baumann 2000). As shown in Figure 4-1, a decrease in cancer at
the site was noted immediately after the plant closure. An increase in
cancer was also noted immediately after dredging and was probably due
to the exposure of fish and their prey to higher concentrations of PAHs
in sediment and water during dredging. Later sampling, however,
showed decreases in cancer, suggesting that the increase during dredging was a short-term phenomena. Within 5 years after remediation, the
cancer incidence was lower than the pre-dredging data, presumably as a
result of the dredging. However, it is unclear to what extent continued
natural attenuation, as evidenced by the reduction in observed cancer
after plant closure but before dredging, could have reduced cancer incidence in the same time frame.
Marathon Battery, New York
The ability to achieve remedial action objectives and long-term risk
reduction with dredging was demonstrated at Foundry Cove of the
Marathon Battery, NY, site. Foundry Cove is a small body of water adjacent to the Hudson River about 85 km north of New York City. The
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60%
Cancer
% of Fish
50%
40%
48%
Pre-cancerous Lesions
46%
38%
30%
23%
22%
20%
18%
10%
10%
11%
7%
0%
0%
1982
1987
1992
1993
1998
Year
FIGURE 4-1 Prevalence of cancer and pre-cancerous lesions in Black River bullheads before and after dredging. The primary source of contamination was shut
down in 1983, and dredging occurred in 1989-1990. Source: Adapted from
Baumann 2000.
Marathon Battery Company discharged cadmium, nickel, and cobalt
during the manufacture of batteries through the plant’s outfalls, located
beneath the Cold Spring pier and in the East Foundry Cove Marsh.
About 50 metric tons of nickel-cadmium waste is estimated to have been
discharged from 1953 to 1971 (Levinton et al. 2006).
The site comprises six separate regions. West Foundry Cove borders the eastern shore of the Hudson River and is connected to East
Foundry Cove via an opening in a railroad trestle. The most contaminated sites are East Foundry Cove Marsh (13 acres), East Foundry Cove
(36 acres), East Foundry Pond (3 ac), and the Cold Spring Pier area (~5
acres) that borders the Hudson River to the north. Constitution Marsh
(281 acres) is to the south and is less contaminated (see Figure 4-2). As
summarized in the record of decision and summary to the committee,
core sediment samples collected from East Foundry Cove during the remedial investigation ranged from 0.29 to 2700 mg/kg cadmium and had
100
FIGURE 4-2 Map showing the Marathon Battery Superfund site on the Hudson
River, NY adjacent to Cold Spring, NY. Dredging was conducted in the Cold
Spring Pier area, East Foundry Cove, and East Foundry Pond. East Foundry
Cove Marsh was excavated and capped; Constitution Marsh was not remediated.
Source: EPA 2006a (Marathon Battery Superfund Site, May 10, 2006).
a mean of 179.3 mg/kg (median = 5.6 mg/kg).4 Samples collected in the
Pier area (a much larger area than what was actually dredged) ranged
from 1.2 to 1,030 mg/kg for cadmium and had a mean of 12.6 mg/kg
These data are for all depths. The mean for each sampled depth is 439.4
mg/kg (0-10 cm), 50.5 mg/kg (10-25 cm), and 2.1 mg/kg (25-50 cm). This sampling
was apparently conducted in 1984 by Acres in support of the remedial investigation (EPA 1989b). These summary statistics from EPA do not include other sampling data collected in 1989 (USACE 1992).
4
101
(median = 3.9 mg/kg)5 (EPA 1989a; 2006a [Marathon Battery Superfund
Site, May 10, 2006]). A human health risk assessment that considered fish
and crab consumption and exposure to suspended sediment in water
concluded that achievement of a sediment cadmium concentration of 220
mg/kg would be protective. Sediment bioassays indicated that 10-255
mg/kg would be protective of aquatic life. A cleanup level of 10 mg/kg
was set and believed to be achievable by removing the top 1 ft of sediment (EPA 1989a). About 80,000 cy of sediment was dredged from the
contaminated areas of East Foundry Cove, East Foundry Cove Pond, and
the Cold Spring Pier area from 1993 to 1995. In contrast, East Foundry
Cove Marsh was dry excavated to a 100 mg/kg limit and then capped
with a bentonite and geotextile blanket followed by 1 ft of sandy marsh
planting material. No active remediation was implemented in the Constitution Marsh6 or the West Foundry Cove (EPA 1995; 2006a [Marathon
Battery Superfund Site, May 10, 2006]).
Post-dredging verification sampling was conducted to establish
whether cleanup levels had been met. EPA states “In the Hudson River
and East Foundry Cove, an average of 10 mg/kg cadmium remained,
which was consistent with the ROD requirement that at least one foot of
sediment and 95% of the contamination be removed” (EPA 1998a). The
record of decision also required long-term monitoring at the site for
thirty years after completion of the remedial action (AGC 2001). Figure 43 presents median cadmium concentrations7 at the East Foundry Cove
portion of the site before and after remediation from this monitoring
program. These data indicate that surficial sediment concentrations were
reduced as a result of dredging, and the concentrations have not re-
These data are for all depths. The ROD describes these data as being from 85
locations covering 465 acres with cores down to a depth of 137 cm. This is a
greater area than the approximate 5 acres that was dredged.
6 “Although cadmium-contaminated sediment hot spots were identified in
Constitution Marsh, to remediate these sediments would have had a significant
adverse impact on the marsh’s sensitive ecosystem. In addition, the cadmiumcontaminated sediments would eventually be covered with clean sediments following the remediation of the cadmium contaminated sediments in East Foundry Cove marsh. Therefore, long-term monitoring was selected for Constitution
Marsh” (EPA 1995).
7For information on the distribution of data, see Figure 4-4.
5
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Nov-04
Nov-03
Oct-02
Apr-01
Apr-00
Apr-99
Apr-98
Apr-97
Jun-96
Mar-96
Nov-95
Post-Remed.
200
180
160
140
120
100
80
60
40
20
0
Pre-Remed.
Cd Concentration (mg/kg)
EFC - Median Sediment Cd Concentrations
FIGURE 4-3 Median sediment cadmium concentrations, East Foundry Cove,
Marathon Battery Site. Dredging occurred in the fall of 1993 and between April
1994 and February 1995. Data are the median of cadmium concentrations from
five locations within the dredged area of East Foundry Cove. Pre-remediation
data are from “samples obtained by Malcolm Pirnie and others prior to the remedial action. These are the reported data closest to the present LTM [long-term
monitoring] sampling locations.”8 Post-remediation data (presumably confirmation sampling immediately following dredging) are the “average value of either
the two closest post-remediation sample node locations or the analytical results
of the various testing agencies for the same node location.” The remaining data
(with dates specified) are from the long-term monitoring (EPA 2006a [Marathon
Battery Superfund Site, May 10, 2006]; the origin of the data table from which
this figure was constructed was not noted in the summary from EPA).
turned to pre-dredging levels. Note that although cleanup levels were
achieved immediately after dredging, median concentrations have since
fluctuated above and below the 10 mg/kg cleanup level. Figure 4-4 presents the frequency of occurrence of sediment cadmium concentrations
in the period 1995-2000. The figure shows a reduction in surficial concentration and also shows that many individual sediment samples show
concentrations greater than 100 mg/kg (log10 = 2). Concentrations in some
sediment samples are indistinguishable from the original sediment concentration distribution. Independent sampling and analysis by Mackie et
It is unclear exactly when these data were collected and to what sediment
depth. According to the ROD, East Foundry Cove sediment samples were collected by Acres in 1985, Ebasco in 1988, and Malcolm Pirnie in 1989.
8
103
al. (2007) had similar results. Their 2005 sampling of the dredged area of
East Foundry Cove showed a median cadmium concentration of 39.2
mg/kg in the top 5 cm of sediment (16 samples ranging from 2.4 to 230.4
mg/kg, mean 59.7, SE 16.8).
The significance of occasional high residual concentrations after
dredging can be evaluated by examination of ecologic exposure before
and after dredging. Figure 4-5 is a summary of long-term monitoring
data collected for 5 years post-dredging (AGC 2001) and shows the ratio
of pre-remediation to post-remediation tissue concentrations in various
plants and animals. The data most relevant to the dredging (which occurred only in East Foundry Cove, East Foundry Pond, and the Cold
Spring Pier area) are the water chestnut in East Foundry Cove, which
show improvement and the benthic invertebrates in East Foundry Cove,
which show an increase after remediation and then a decrease to preremediation levels after 5 years.9 Bioaccumulation studies (using in-situ
Pre-Remed.
East Foundry Cove
Log10 Cd Concentration, mg/kg
Post-Remed.
4
Nov-95
Mar-96
3
Jun-96
2
Apr-97
Apr-98
1
Apr-99
0
Apr-00
Apr-01
-1
-2
-1.5
-1
-0.5
0
0.5
z score
1
1.5
2
Oct-02
Nov-03
Nov-04
FIGURE 4-4 Log probability plot of sediment cadmium concentrations, East
Foundry Cove, Marathon Battery. The z score indicates the distance from the
mean in standard deviations; z scores greater than 0 exceed the mean. The source
of these data is described in Figure 4-3.
Because birds are highly mobile, contaminant exposures to wood ducks,
swallows, and marsh wrens cannot be directly linked to the dredged areas. Also,
Constitution Marsh did not undergo active remediation; East Foundry Cove
Marsh was excavated and capped.
9
104
100.00
1996
1997
1998
1999
2000
10.00
1.00
0.10
BENTHIC
INVERTEBRATES
(EFC)
WATER
CHESTNUT (CM)
WATER
CHESTNUT (EFC)
CATTAILS (EFCM)
CATTAILS (CM)
MARSH WREN
(whole body)
SWALLOWS
(whole body)
WOOD DUCK
(kidney)
0.01
WOOD DUCK
(liver)
Ratio to Historical Concentrations
enclosures) were also conducted at the site. Data from East Foundry
Cove and the Cold Spring Pier area generally indicate declines in accumulated cadmium body burdens compared with pre-remediation values
(AGC 2001; EPA 2003b).
The data suggest that in at least some cases dredging can achieve
and maintain reductions in sediment concentrations and body burdens
of contaminants although occasional measurements of elevated concentrations complicate the interpretation of the results. In addition, there
may be short-term increases in body burdens in species directly affected
by the remediation (such as the benthic organism data). The fact that individual sediment samples exhibited elevated concentrations emphasizes that evaluation of the performance of any remedy requires adequate monitoring of key indicators before and after remediation to fully
characterize the distribution of concentrations so decisions are not driven
by low probability events.
FIGURE 4-5 Ratio of cadmium body burdens in various animals and plants after
dredging to pre-dredging levels for the period 1996-2000, Marathon Battery Site,
NY. The data are the mean value except for the benthic invertebrates for which
only one sample was collected. (The benthic invertebrate samples were “a combination of oligochaete worms and chironomid midge larvae, with a minimum
aggregate weight of one gram.”) The number of samples in each mean varies.
CM = Constitution Marsh; EFCM = East Foundry Cove Marsh; EFC = East Foundry Cove. Note that only the water chestnuts (EFC) and the benthic invertebrates
(EFC) were from locations where dredging alone was conducted. Source: Data
from AGC 2001.
105
Summary
Dredging remains one of the few options available for the remediation of contaminated sediments and should be considered, along with
other options, for managing the risks that they pose. The need to remove
contaminated sediments can be particularly acute at sites where navigational channels need to be maintained or where buried contaminated
sediment deposits are likely to be subjected to erosion and transport
from high flows or changes in hydrologic conditions. As shown at the
Grasse River and other sites, dredging can achieve removal of sediments
and much of the contaminants they contain. Mass removal alone, however, may not achieve risk-based goals, which should be the basis for
remedy selection.
The results at the Marathon Battery and other sites outlined above
show that under some conditions dredging can achieve cleanup levels
and aid recovery of biota at contaminated sediment sites. As indicated
previously, it is often difficult to evaluate the effectiveness of dredging
alone as a result of reductions in ongoing sources of contamination (for
example, outfalls and atmospheric deposition), the use of combination
remedies, and natural burial of existing contamination. As illustrated in
the examples above, measurement of time trends in sediment and water
contaminant concentrations prior to and after dredging can help identify
changes due to dredging as well as evaluate the risks of not dredging.
Reference sites are also useful in contrasting the contaminant dynamics
and risk reduction due to dredging with that caused by these other processes. To assist in evaluating the performance of dredging, it is important
to monitor a range of dredging performance metrics linked to risk reduction, as was done at Marathon Battery.
FACTORS AFFECTING DREDGING EFFECTIVENESS
A variety of factors and site characteristics influence dredging effectiveness and can limit the ability to achieve cleanup levels and remedial action objectives. However, it is generally not possible to definitively
identify the specific conditions or factors that determined success or failure in a particular project. The committee strived to identify the condi-
106
tion or conditions that appeared to significantly contribute to the project’s outcome and summarized those herein.
Foremost among the factors that influence dredging effectiveness
are whether a site has been adequately characterized and whether ongoing sources of contamination have been controlled. Site characterization
and source control require a firm understanding of the nature and distribution of the contamination, any potential sources that contribute to the
contamination burden in the watershed, and the processes influencing
site risks and their attenuation. A strong understanding of the extent of
contaminants in place and the contribution of outside sources is essential
to developing an effective conceptual site model and remedial plan to
eliminate or lessen contaminant exposures and risk. The influence of
source control and site characterization on remedy effectiveness will be
illustrated through experience at particular sites in later sections. These
factors, however, influence the success of all sediment remediation efforts, not just dredging.
Destruction of the benthic community and removal of habitat is
unavoidable with all dredging projects and represents an immediate
negative effect to the existing benthic community. This effect also occurs
with other active remedial efforts such as capping. As such, the ecologic
benefit of the current habitat needs to be an important part of the decisions in determining whether or not to dredge. In a net risk reduction
framework, the habitat destruction will be compared to the benefits of
removing contaminated sediment, bearing in mind the post-dredging
substrate’s desirability as a habitat (or the substrate created following
backfilling). 10 Recovery after disturbance is typically relatively rapid
with estimates of benthic recovery rates ranging from several months to
several years (Qian et al. 2003; USACE 2005). Immediately after destruction of the habitat, hardy, opportunistic organisms such as polychaetes
(oligochaetes in freshwater) and small bivalves colonize surficial sediments. Subsequently the population increases in diversity and abun-
As described by EPA (2005a): “While a project may be designed to minimize
habitat loss, or even enhance habitat, sediment removal and disposal do alter the
environment. It is important to determine whether the loss of a contaminated
habitat is a greater impact than the benefit of providing a new, modified but less
contaminated habitat.”
10
107
dance. Recovery occurs when the site returns to pre-disturbance conditions or does not differ significantly from a reference area.
In contrast to the above factors which affect many or all remedial
approaches, dredging projects are specifically influenced by two additional processes that weigh heavily on the effectiveness of dredging:
• Resuspension of sediments and release of contamination during
dredging.
• Generation of residual contamination giving rise to potential
long term exposure after dredging.
As described in Chapter 2, resuspension, release, and residuals occur to some extent with all dredging projects. Resuspension refers to
sediment that is disturbed during dredging and transported out of the
dredging area. Exposure to resuspended sediments is generally transitory and ends soon after the completion of dredging. Residuals are the
contaminated sediments exposed after conclusion of the dredging and
can lead to longer term exposure and risks to organisms. Release of contaminants to the liquid phase (for example, solubilized PCBs) can occur
from both resuspended and residual sediments. In the case of strongly
sorbing contaminants, it is often assumed that the fraction of sediment
resuspended corresponds to the fraction of contamination released and
transported down current. Sediment resuspension and contaminant release, however, may not be closely related if there are large dissolved or
separate phase releases or if release from residuals is substantial.
Figure 4-6 summarizes the amounts of resuspension and residual
contamination that have been observed in a variety of dredging projects
(Patmont 2006). The sediment resuspension data points are the fraction
of sediment resuspended during dredging. (The fraction is the mass of
suspended solids measured at some distance downstream of the dredge
[typically less than 100 ft] divided by the mass dredged on a dry weight
basis.) These data are from a variety of sources including consultant reports and the open literature. As such, sampling methods and approaches used to estimate these fractions can vary depending on the objectives of the study, the nature of the project, and site conditions. The
residual data in the figure are the fraction of contaminant mass (not
sediment mass) remaining post-dredging compared to the estimated
108
FIGURE 4-6 Sediment or contaminant mass resuspended or left as a residual.
Resuspension data (squares, left) are from a variety of projects gleaned from the
literature. Residuals data (diamonds, right) are from dredging at 12 projects (average refers to the median in the figures). Source: Patmont 2006. Reprinted with
permission; copyright 2006, Anchor Environmental LLC.
contaminant mass removed in the last production dredging cut at 11
sites (residual fractions are determined based on a chemical massbalance approach [see Patmont and Palermo 2007]). The projects span a
range of physical settings, operating conditions, and data collection
methods. Despite variations among sources of the resuspension and residuals data, the distribution shown is a useful compendium of existing
data and likely to indicate at least the magnitude of expected residuals
and resuspension rates.
The figure suggests that about half the dredging projects have resulted in resuspension that amounts to 1% or less of the mass of sediments dredged.11 About half the dredging projects have resulted in a residual contaminant mass that amounts to 5% or less. Although the
resuspension losses and residuals are small relative to the total mass
It should be noted that the resuspended solids fraction measured downstream likely underestimates the total contaminants in the water column because
some dissolved releases are likely to occur from solids that remain within the
dredge footprint, that is, freshly exposed and redeposited sediments.
11
109
dredged, the availability of the contaminants to organisms may be
higher than prior to dredging because of exposure to contaminants in the
water column (resuspension) or at the sediment surface (residual). The
residuals may be especially problematic in that the concentration in the
residual is similar to the average concentration in the dredged sediment
(Reible et al. 2003; Patmont 2006) and directly accessible to organisms
that live at or interact with the sediment-water interface. Because of the
presence of the residuals, surface concentrations may not change or may
even increase when compared with pre-dredging conditions.
Patmont and Palermo (2007) used a similar (though not identical)
dataset to investigate the influence of site-specific factors on residual
contamination after dredging. In this analysis, they concluded that
Similar generated residual percentages were observed for both mechanical and hydraulic dredges. The available data suggest that
multiple sources contribute to generated residuals, including resuspension, sloughing, and other factors. However, on a mass basis, sediment resuspension from the dredgehead appears to explain
only a portion of the observed generated residuals, suggesting that
other sources such as cut slope failure/sloughing are likely quantitatively more important. The available mass balance data also indicate that the presence of hardpan/bedrock, debris, and relatively
low dry density sediment results in higher generated residuals.
Figure 4-7, from Patmont and Palermo, shows the influence of debris and/or hardpan and sediment bulk density on the estimated residual
at 11 dredging project sites. (Figure 4-7 has one less case study than Figure 4-6.) Higher amounts of debris, the presence of hardpan, and low
sediment bulk density all contribute to higher generated residuals.
An examination of dredging at the various sites included in Appendix C indicated that resuspension, release, and residuals can all be
influenced by site conditions such as those discussed by Patmont and
Palermo (2007) and by the manner in which dredging is implemented.
The next section discusses the role that each of those may play in limiting
dredging effectiveness on the basis of experience at specific sites where
dredging was used; taken together, the experience illustrates site-specific
conditions or activities that contribute to or limit dredging effectiveness.
110
FIGURE 4-7 Estimated generated residuals for 11 projects with data broken
down on whether debris, rock, or hardpan was prevalent at the site. Source:
Patmont and Palermo 2007. Reprinted with permission from the authors; copyright 2007, Battelle Press.
The objective is to identify conditions under which dredging might be
effectively implemented and conditions that could discourage the use of
dredging because of its inability to meet desired cleanup levels or remedial action objectives.
Resuspension and Release of Dredged Contaminants During Removal
Resuspension and release of contaminants during dredging are
among the most important adverse effects of dredging. As shown in Figure 4-6, up to 10% of the mass of sediment dredged can be resuspended
during dredging. Early illustrations of the potential for dredging to give
rise to at least short-term increases in adverse effects on organisms can
be found in the previously discussed Black River, OH, and Marathon
Battery, NY, sites. Those increases were probably due to exposure to
more highly contaminated sediment or resuspension of sediment and
contaminants during dredging. In the next two sections, the effect of
sediment resuspension during dredging and the duration of those effects
will be illustrated by experiences at other sites.
111
Effects of Resuspension and Release
Grasse River, New York
Monitoring of dissolved PCB concentrations in the water column
during the 1995 non-time-critical removal in the Grasse River, Massena,
NY indicated water concentrations above the 2 μg/L PCB action level at
several time points and as high as 13.3 μg/L PCBs (BBL 1995).12 These
concentrations were detected along the perimeter of the project, beyond
three lines of silt curtains that were used in an effort to contain suspended sediment. The interior curtain was extended to the bottom, and,
as stated in the documentation report (BBL 1995), “the lowering of both
the boulder and the inner, secondary curtains to the River bottom greatly
enhanced TSS [total suspended solids] containment.” This site also used
caged fish before, during, and after dredging to monitor bioaccumulation of PCBs (BBL 1995). The study concluded that the increases in caged
fish exposed for six weeks during dredging had increases in PCB concentrations 20 to 50 times higher than those observed in the pre-dredging
time frame and increases of that magnitude suggest that uptake of PCBs
was affected by the release of PCBs to the water column during dredging.
However, the report also states that some of the increases in caged fish
may be attributable to higher water temperatures during the dredging
exposures.13
12 In contrast, site-wide sampling the year prior to the dredging generally
showed no quantifiable PCB concentrations throughout the river (that is, concentrations were less than 0.5 or 0.7 μg/L depending on the Arochlor) (BBL 1995).
13The analysis is further complicated because the increases in fish upstream of
the removal were proportionally greater than those downstream. For the 6-week
exposures during dredging, lipid-normalized PCB concentrations at the two upstream cages were on average 59 times higher than pre-dredging exposures,
while downstream cage locations were on average 35 times higher. However, on
an absolute basis, the increases were less pronounced upstream (average increase
of 658 ppm PCBs on a lipid basis) relative to the increases observed at the downstream locations (average increase of 2,394 ppm PCBs on a lipid basis). It is suspected that the greater proportional increases in the upstream cages (150 ft upstream of the removal area) resulted because the pre-dredging concentrations in
the upstream cages were lower and, thus, increases in PCBs at both locations will
result in a greater proportional increase at the upstream cage locations and be-
112
A second dredging project took place in the Grasse River in 2005.
This demonstration project was intended to remove approximately
64,000 cy of PCB-contaminated sediment from 3 work zones, but ultimately removed about 24,600 cy of sediments from approximately 40%
of the targeted area. Water column sampling during this demonstration
project showed about one-fourth of the 100-odd measurements taken
over the course of the project exceeded the 2 μg/L PCB action level in the
most downstream sampling site adjacent to the silt curtain (EPA, unpublished data, 4/18/2006). All measured water column concentrations
throughout the river prior to the demonstration dredging project were
less than 0.065 μg/L total PCBs.
As shown in Figure 4-8, increases in water-column PCB concentration were noted downstream of the dredging operation. TSS increased in
concert with PCBs at the sampling locations adjacent to the site, but
downstream PCB concentrations remain high when TSS decreased back
to upstream values. The presence of dissolved PCBs that would not settle
may account for the different behavior downstream. Because dredging is
a technology designed to remove solids, dissolved contaminants are contained much less effectively. Similar behavior would be expected for remedial dredging of sediments containing fluid contaminants, such as
nonaqueous-phase liquids. In addition, operational controls on resuspension of sediment particles are expected to have less effect on dissolved-phase or fluid-phase contaminants. Silt curtains, for example, are
designed to provide additional residence time to encourage particle settling and would have less influence on non-settleable contaminants.
The increase in suspended solids and their flux down river was not
associated with high-flow events and therefore was likely due to dredging-related resuspension processes. A horizontal auger dredge was used
at this site for most of the demonstration dredging. A horizontal auger of
the size used is limited in its ability to dredge effectively in the presence
of stone or debris 4 in. or larger. A separate debris-removal operation
was implemented to eliminate larger stones and debris. As is typical, an
open bucket was used to allow sediment captured with the debris and
stone to redeposit on the bottom. Such operations can increase the entry
cause low flows during the removal created conditions favorable for the upstream movement of water (J. Quadrini, written commun., March 23, 2007).
113
FIGURE 4-8 Average water-column PCBs and TSS during debris and sediment
removal activities (conducted simultaneously) in the Grasse River during 2005
demonstration dredging program. ROPS-WCT5 is a sampling site upstream of
dredging area; sampling locations D1 through D4 are adjacent to the dredging
site in an upstream (D1) to downstream (D4) direction; ROPS-WCT14 and ROPSWC131 sites are about 0.5 and 1.0 miles downstream of the dredging activity,
respectively. Error bars represent ±2 standard errors of the mean. Source: Connolly et al. 2006. Reprinted with permission from the authors; copyright 2006,
Quantitative Environmental Analysis.
of suspended solids into the water column and thus increase the contaminant burden in the water column. Similar problems occur in the use
of clamshell-bucket dredging of sites laden with debris. The occasional
inability to close the bucket completely because of debris interference can
increase resuspended solids and thus resuspension of contaminants.
A more detailed depiction of the PCB concentrations in water seen
approximately 0.5 miles downstream of the dredge site is presented in
Figure 4-9. The pre-dredging baseline PCB concentrations are low; during dredging activities, these concentrations generally increase and occa-
114
sionally exceed the 2 μg/L PCB action level; after dredging, concentrations decrease back to baseline concentrations (see Figure 4-9). It is estimated that during dredging activities about 3% of the PCBs removed
from the river bottom were released downriver, largely as PCBs that had
desorbed from resuspended sediments (Connolly et al. 2007).
The overall effect of resuspension and release during the dredging
operation can be seen more clearly by examining PCB concentrations in
fish at the Grasse River site. Figure 4-10 shows PCB concentrations in
spottail shiner measured every fall from 1998 to 2006.
Spottail shiners are useful indicators because they forage only over
a limited area (Becker 1983) and, being small, respond quickly to increases in PCB concentrations (Connolly et al. 2006). During 2005, the
fish sampling coincided with dredging activities and PCB concentrations
in the spottail shiner increased dramatically. The following year, PCB
concentrations in shiners decreased to levels seen prior to dredging.
There was a statistically significant increase in the downstream locations,
but there is insufficient information to evaluate trends associated with
dredging, because only a single post-dredging monitoring period was
ROPS-WCT14
Characteristic
Baseline
FIGURE 4-9 PCB concentrations in water samples collected approximately 0.5
miles downstream of the dredging operations in the Grasse River (NY). Dredging began approximately June 8, 2005 and ended October 21, 2005. Source: Connolly et al. 2006. Reprinted with permission from the authors; copyright 2006,
Quantitative Environmental Analysis.
2006
2003
2004
2005
2001
2002
1998
1999
2000
2006
1998
1999
2000
2001
2002
2003
2004
2005
2006
2005
2004
2003
2001
2002
2000
1998
1999
2006
2000
2001
2002
2003
2004
2005
1998
1999
FIGURE 4-10 Lipid-normalized PCB concentration in resident spottail shiners in Grasse River, Massena, NY. Locations are from
upstream (Background and Near Outfall 001) to downstream (Near Unnamed Tributary [about 2 miles downstream] and Mouth
of River [about 6 miles downstream]). Error bars represent ±2 standard errors of the mean. Dredging at the site took place June to
October, 2005; the 2005 fish tissue sampling was in September, 2005. Source: L. McShea, Alcoa Inc., unpublished material, February 27, 2007.
115
116
available for PCB body-burden analysis (see Box 4-1). Additional data
collection and detailed analyses of the Grasse River data are ongoing by
EPA and Alcoa.
Duwamish Diagonal, Washington
In the Duwamish/Diagonal CSO, WA, early-action sediment dredging project, high PCB concentrations (above the pre-dredging surface
concentrations) were found in sediments in and outside the dredge
prism during a pre- and post-dredging sampling program (EcoChem Inc.
2005). During dredging, several complaints were logged about poor
dredging practices that may have contributed to resuspension and release of contaminants. The dredged areas were capped, and to address
the unexpected contamination created outside the dredge prism, 6-9 in.
of sand was added to areas adjacent to the dredging area. Water quality
monitoring during dredging indicated that turbidity standards were exceeded on several occasions (particularly in the first 2 weeks). Total PCBs
and dissolved mercury (measured only during the first 8 days of dredging) were below water quality standards even at the highest turbidity
values (EcoChem Inc. 2005).
As part of the overall RI/FS of the Lower Duwamish River Superfund site and other studies, fish tissue samples were collected both before and after the Duwamish/Diagonal dredging project. These data suggest that fish-tissue PCB concentrations were greater after dredging
activities (Figure 4-11). While there is only one data point for the targeted
species collected several years prior to dredging, other fish data collected
in the project area in the years prior to the early action (that is, between
1992 and 2000) indicate that tissue levels remained steady during this
period (J. Stern, King County Department of Natural Resources and
Parks, personal commun., April 20, 2007). The exposure dynamics of
resident fish are complex, so monitoring data alone are unable to directly
implicate dredging as the cause of the apparent “spike” in tissue PCB
concentrations. However, the timing of the increased fish concentrations,
the rapid decrease after dredging, and corroborating fish bioaccumulation modeling (Patmont 2006; Stern and Patmont 2006; Stern et al. 2007)
are suggestive of a dredging-related release.
117
BOX 4-1 Statistical Analysis of Fish PCB Body Burdens,
Data
The data analyzed for this report (EPA, unpublished data, 4/18/2005) included
percent lipid and total PCB concentrations in fish tissue for smallmouth bass (fillets),
brown bullhead (fillets), and spottail shiner (whole body) sampled from the years
1993-2004 (pre-dredging), 2005 (during dredging) and 2006 (post-dredging). Brown
bullheads and smallmouth bass were sampled from 4 areas: Background (upstream
of dredging) and the Upper, Middle, and Lower stretch of the river (increasing distance downstream from dredging). Spottail shiners were sampled from the Background, Near Outfall 001, Near Unnamed Tributary, River Mouth areas (see Figure 410 legend for spottail shiner location details).
Methods
Temporal trends in fish tissue PCB concentrations and region specific effects
were established based on linear regression models using monitoring year, percent
lipid content, and sampling region as independent variables. PCB concentrations
were centered upon their sampling region mean. The Box-Cox transformation (Box
and Cox 1964) was parameterized in the regression model likelihoods to allow the
data to optimally choose possible transformations. Analyses were stratified by fish
species. Nonlinear trends in time were considered (Stow et al. 1995), the results of
which led to interpretations that were qualitatively similar.
Different detection limits were reported for these data. Regression model inference was based on maximum likelihood treating the below detection limit data as left
censored (Helsel 2005). (Data are left censored if their numeric value only indicates
they are less than some given threshold such as the detection limit.) This method has
been suggested as an alternative to substitution based techniques such as replacing
non-detects with half their detection limit.
Results
Fish tissue PCB concentrations were transformed using natural logarithm for
all analyses. Lipid adjusted fish tissue PCBs sampled in the Background region were
significantly lower than those sampled in other regions, smallmouth bass (p<0.01),
brown bullhead (p<0.01), spottail shiner (p<0.01). Region specific significant decreasing temporal trends based on pre-dredging data (1993-2004) were established for all
species.
118
BOX 4-1 Continued
We explored whether PCB concentrations during dredging (2005) were significantly greater than that expected from the temporal trend established on the predredging data (1993-2004). The post-dredging (2006) data were also compared to the
established pre-dredging temporal trend.
For the brown bullhead, the during dredging lipid adjusted average PCB concentrations were larger than the upper 95% limit of the time trend prediction interval
for all regions. Results for the Background stretch region is interpreted with caution
due to small sample size. For the smallmouth bass, the during dredging lipid adjusted average PCB concentrations were larger than the upper 95% limit of the time
trend prediction interval for all regions, except for the Background stretch. For the
spottail shiner, the during dredging lipid adjusted average PCB concentrations were
larger than the upper 95% limit of the time trend prediction interval for “Mouth of
river” and “Near unnamed tributary” areas. Concentrations during dredging in the
background and “Near outfall” stretches were within these time trend based prediction limits. Due to small sample sizes results for spottail shiners are interpreted with
caution.
The post-dredging (2006) lipid adjusted average PCB concentrations decreased
significantly compared to those measured during dredging (2005) for all fish species
and all sites, except the background regions for smallmouth bass and brown bullhead. The lipid adjusted fish tissue PCB concentrations post dredging (2006) were
within the range (95% prediction intervals) predicted by the temporal trend established on the pre-dredging data (1993-2004) for all species and regions except for
smallmouth bass in the lower and middle regions which remained above the 95%
prediction intervals.
Remarks
Additional post-dredging sampling points will be needed to evaluate longterm dredging effectiveness. Longitudinal monitoring (encompassing pre- and postdredging time frames) can be used to statistically compare trends in contaminant
concentrations before and after dredging and better associate changes with dredging.
Fox River, Wisconsin
A further illustration of the influence of sediment resuspension on
dredging effectiveness can be found in various demonstration projects
conducted at the Fox River, WI. At Deposit N, silt curtains were used in
1998 to contain any resuspended sediment, and downstream turbidity
119
4500
4000
English
EnglishSole
Sole
Shiner
ShinerSurfperch
Surfperch
Dredging and Backfilling
ug/kg ww)
PCBs in Fish (whole body µg/kg
5000
3500
3000
2500
2000
1500
1000
500
0
Jul-96 Jul-97 Jul-98 Jul-99 Jul-00 Jul-01 Jul-02 Jul-03 Jul-04 Jul-05 Jul-06
Time
FIGURE 4-11 PCB concentrations in fish collected at the Duwamish Diagonal site
pre- and post-dredging. Two trophic levels of fish (English sole and shiner surfperch) of consistent size range were collected within 1 km of the site at the time
points indicated on the figure. Error bars represent the minimum and maximum
values. Dredging was conducted at the site between November 14, 2003 and
January 20, 2004. Capping was completed by February 29, 2004. Sources:
Adapted from Patmont 2006; Stern and Patmont 2006; Stern et al. 2007. Fish data
from King County 1999; King County, unpublished data, 2007; Windward 2005,
2006.
was found to be no greater than that at upstream sampling locations.
Upstream and downstream turbidity levels were also very similar in
1999, when silt curtains were not used (Foth and Van Dyke, 2000). Nevertheless, during 1998 dredging, water-column PCB concentrations averaged 11 ng/L downstream of dredging and 3.2 ng/L upstream. Prior to
dredging, the average PCB water column concentrations were similar
upstream and downstream of the dredging area. Similar increases (24
ng/L downstream vs 14 ng/L upstream) were observed in 1999. Overall,
about 4% of the PCB mass removed from the deposit was released to the
water column by dredging (Malcom-Pirnie and TAMS, 2004). The demonstration conducted in Sediment Management Units 56 and 57 (SMU
56/57) in 1999 and 2000 targeted a deposit of about 80,000 cy near the
outfall of a recycling mill in Operable Unit (OU) 4, a relatively lowenergy estuarine reach of the river. During dredging in 1999, in which
120
31,000 cy were removed, about 2.2% of the PCBs dredged were estimated
to have been resuspended and transported downstream (Steuer 2000).
Direct estimates of contaminant release can rarely be used to guide
day to day dredging operations because water column samples may take
several days to analyze. Suspended-solids measurements are sometimes
used to provide real-time feedback for the dredging operation. As indicated in Box 4-2, however, the correlation between suspended solids and
even strongly sorbing contaminants, such as PCBs, may not be adequate
to guide operations appropriately.
BOX 4-2 Correlations between Suspended Solids and
Contaminant Concentrations
Although resuspension of sediment is largely viewed as the source of contaminant losses, any contaminant that partitions rapidly from the sediment to
the water column will quickly cease to be related to resuspended-sediment concentrations. As shown by the Grasse River 2005 data (Figure 4-8), the watercolumn concentrations did not generally correlate with suspended-solids concentrations. Turbidity is typically used as a rapid surrogate measure of suspended solids and can be useful to indicate suspended solids if a site-specific
correlation between the two quantities can be found. As suggested in the text,
however, turbidity and contaminant concentration might not correlate. In an
effort to understand the nature of the PCB releases from the Grasse River site,
samples were taken from adjacent areas in and outside the silt-containment device surrounding the dredge area. Analysis indicated that although TSS concentrations were about 2 times greater inside the curtain, the dissolved-PCB concentrations were the same (Connolly et al. 2006). At the downstream sampling site,
about 0.5 mile from the dredging area, about 75% of the PCBs was dissolved
(this is operationally defined as passing a 0.45-μm filter).
At the Fox River SMU 56/57 sites, dredging-related releases resulted in an
increase in downstream dissolved-PCB concentrations of about 59%. However,
little or no difference in turbidity or TSS concentrations between upstream and
downstream locations was detected during dredging, and turbidity and TSS did
not correlate with water-column PCB concentrations (Steuer 2000). Those results
indicate that turbidity and TSS were of little value as indicators of water-column
PCB release. Dissolved and fluid contaminants, such as nonaqueous-phase liquids, will not be well characterized by TSS or turbidity monitoring. A good correlation between suspended solids and contaminant resuspension would be expected if the contaminant remained strongly associated with the solid phase.
121
Duration of Effects from Resuspension and Release
Increases in contaminant concentrations in the water column and
fish during or immediately after dredging may be short-lived. As shown
in the Black River, cancer prevalence at the site increased following
remediation, but then declined dramatically shortly afterward (Figure 41) (Baumann 2000). At Marathon Battery, cadmium concentrations in
benthic invertebrates increased following dredging, before declining
again (Figure 4-5). In addition, PCB concentrations in the fish in the
Grasse River declined substantially a year after concentrations spiked
during dredging in 2005 (Figure 4-10).
There is also evidence that contaminant releases from dredging can
be reduced. At the GM Massena project in the St. Lawrence River, a
sheet-pile wall was erected around the dredging zone because silt curtains were unable to withstand the river currents. The interlocking steel
sheet piling enclosed the area to reduce offsite migration of sediment
(EPA 2005a). Sheet piling was possible, however, because isolation was
required for only a small portion of the river bottom. During dredging,
turbidity and PCB and PAH concentrations were measured downstream
of the sheet piling to monitor for potential releases into the river. PAH
results were consistently below detection limits, and sampling ceased
after 19 days. PCBs were monitored over about 3 months. All samples at
the monitoring stations were well below the action level of 2 μg/L; the
maximum was 0.32 μg/L, and most of the samples were below the detection limit (BBL 1996).14
The examples cited indicate that resuspension and contaminant release during dredging can limit at least short-term dredging effectiveness. Large dredging projects that may continue for years or decades are
more likely to exhibit more serious problems associated with resuspension and contaminant release than the projects of shorter duration that
were examined. For example, dredging could continue for over 25 years
The sheet piling may have increased potential concerns about residual contamination while decreasing resuspension losses. In the sheet piled area with the
most heavily contaminated sediment, PCB concentrations in water were greater
than 20 μg/L a week after dredging stopped. A week later, the concentrations
had declined to below 2 μg/L, and the sheet piling was removed. When the sheet
pile had been removed, water concentrations dropped to near the detection limit.
14
122
at New Bedford Harbor (Dickerson and Brown 2006) and thus any effects of resuspension would be at least of similar duration. There is no
experience with such large projects, although some have been initiated
(for example, New Bedford Harbor, MA, and Fox River, WI) or planned
(for example, Hudson River, NY). The rate of recovery and time to
achieve remedial goals after long-term exposure to remedial dredging
are not known. Careful design of a monitoring program is needed to
separate short-term from longer-term performance of a remedy.
Generation and Exposure of Residual Contamination
Potentially the most serious limitation to dredging effectiveness is
residual contamination that is left after dredging. As described in Chapter 2, there are two general types of residuals: generated residuals, contaminated sediment that is resuspended during dredging and later redeposited; and undisturbed residuals, contaminated sediments found at
the post-dredge sediment surface that have been uncovered but not fully
removed as a result of the dredging operation (Bridges et al. in press). A
portion of the generated residual may be unconsolidated and potentially
more susceptible to transport. As such, this portion may not be accounted for by confirmation sampling conducted to define postdredging residuals, depending upon the timing of that sampling. The
presence of such residuals directly limits the ability to meet cleanup levels and may also reduce or eliminate opportunities to achieve long-term
remedial action objectives. Findings from several of the studied sites on
the extent of residual contamination after dredging are provided below.
Lavaca Bay, Texas
A dredging demonstration project was conducted in August 1998
to evaluate the use of a full-size hydraulic dredge to remove mercurycontaminated sediment near the outfall of a chloro-alkali manufacturing
facility on the northwest shore of Lavaca Bay, TX.15 This hot-spot area
This description refers to “Phase 1” dredging of the treatability study (Alcoa
2000). The second phase of the study targeted a smaller, less-contaminated shallow water area.
15
123
had the highest sediment mercury concentrations in Lavaca Bay, which
has widespread mercury contamination. Six acres of very soft plastic clay
sediment was dredged in 20 days; 2,300 lbs of mercury was removed
with 60,000 to 80,000 cy of sediment, and placed in a confined disposal
facility (Alcoa 2000). Extensive data on sediment contaminant concentrations (including post-dredging residual sediment in the dredge area),
water quality (including TSS, turbidity and mercury concentrations), and
mercury accumulation in caged oysters were collected at the site. Low
TSS and insignificant mercury were mobilized beyond the curtained-off
zone surrounding the dredge unit. Elevated mercury concentrations in
oysters were within the range of those observed in oysters native to
Lavaca Bay (Alcoa 2000).
The demonstration project is informative because of the efforts to
control sediment residuals. Multiple passes of the dredging operation
were conducted with sampling between passes to define the residual
present after each pass. There was a notable increase in residual concentration between passes 2 and 3, as shown in Figure 4-12, apparently reflecting exposure of more highly contaminated sediment. Overall, the
pass-to-pass concentration changes were not statistically significant, although analyses stratified by subarea did reveal some pass-to-pass concentration changes that were statistically significant (see Box 4-3 and Figure 4-12).
Grasse River, Massena, New York
At the previously discussed 1995 non-time-critical removal action
in the Grasse River, Massena, NY, the average PCB concentrations in
surficial sediments (upper 8 in.) were reduced by only 53% despite removal of as much as 98% of the PCB mass from the sediment column
(Thibodeaux and Duckworth 1999). The site contains numerous rocks
and boulders that contributed to residual contamination and contained
high concentrations of PCBs near the bottom of the sediment column
that could not feasibly be dredged, because of underlying bedrock and
glacial till (hardpan) (see Box 4-4).
124
BOX 4-3 Statistical Analysis of Mercury Concentrations in
Surficial Sediment, Lavaca Bay, Texas
Data
Surficial sediment mercury concentrations in Lavaca Bay (Alcoa 2000) before
any dredging and after four sequential dredging passes were analyzed. Sample locations were identified in four subareas: Capa, North Capa, AA, and the Trench Wall.
Methods
Statistical comparisons of surficial mercury concentrations were based on a linear regression model with indicators of dredge pass (before and after dredging passes
1, 2, 3, and 4) as the independent variables. The Box-Cox transformation (Box and
Cox 1964) was parameterized in the regression-model likelihoods to allow possible
transformations to be chosen optimally. Regression-model inference was based on
maximum likelihood. Stratified analyses were performed for different subareas on
the basis of the Kruskal-Wallis rank sum test (Hollander and Wolfe 1973) because the
samples were small.
Results
Figure 4-12 (left) displays mean surficial mercury concentrations prior to
dredging and after each dredging pass and their corresponding 95% confidence intervals on the log base 10 scale. Mean surficial concentrations after each dredging
pass were not significantly different from the means sampled before dredging. Figure
4-12 (right) displays the distribution of surficial mercury concentrations and corresponding sample sizes stratified by subarea and dredge pass. For Capa, mean surficial mercury concentrations before dredging after dredging passes 1 and 2 did not
differ significantly (p = 0.28). For the Trench Wall, mean surficial mercury concentrations increased sequentially after dredging passes 1, 2, and 3, however these changes
were not statistically significant. The mean concentration after dredging pass 4 was
significantly lower than after dredging pass 3 for the Trench Wall (p = 0.04). Before
dredging, mean surficial mercury concentrations differed significantly among the
four subareas (p = 0.06).
Remarks
Exploratory and followup stratified analysis suggested the geographic subareas of Capa, North Capa, AA, and the Trench Wall as a potential source of surficial
mercury variation. However, sample sizes were insufficient to include this spatial
effect in the regression model while considering variation across dredging times.
0
|
|
|
10
|
|
|
25
|
|
|
50
|
|
|
Mercury mg/kg
100
|
|
|
140
|
|
|
-
-
-
-
N
2
13
8
8
12
5
17
15
2
0
6
10
0
0
0
13
13
8
0
0
FIGURE 4-12 Left, mean and 95% confidence intervals of surficial mercury concentrations (mg/kg) (log base 10 scale) before
dredging and after each of four dredging passes. Right, concentration distribution and sample sizes (N) of surficial mercury stratified by geographic region—Capa (CA), North Capa (NC), AA, and the Trench Wall (TW)—and dredging pass (Pre, 1, 2, 3, 4; pre =
pre-dredging).
Pre
1
TW 2
3
4
Pre
1
AA' 2
3
4
Pre
1
NC 2
3
4
Pre
1
CA 2
3
4
Lavaca Bay Surficial Mercury Data
125
126
BOX 4-4 Description of the Substrate Topography of the
Grasse River During Dredging
“The river bottom, we believe, contains boulders and rock outcrops that
account for these features [seen in the side scan sonar images]. Soft sediment is
intermixed with these features. The river was dredged in early 1900s using
equipment and techniques that may have included blasting, which left a bottom
littered with rocks and boulders, perhaps some outcrops and glacial till. Soft
sediment began to settle on top of this bottom beginning in 1958 when the power
canal ceased contributing flow to the river. All the accumulating soft sediment
contains PCBs because PCBs were discharged from the late 1950s on, and this
contaminated sediment fell in and amongst the rocks and boulders and finally
covered them. So what we encountered is a bottom littered with rock debris intermixed with soft sediment below which is glacial till. The [horizontal auger
dredge] captured some material but the productivity was very low because the
auger couldn’t get down in between the rock debris, so I think that was part of
the problem.”
Source: Connolly et al. 2006.
Immediately following dredging at the 2005 demonstration project
at the Grasse River, residual surficial sediment PCB concentrations (0-3
in.) averaged 150 mg/kg (dry weight), compared to a pre-dredging average of 4.1 mg/kg (Connolly et al. 2006). The increase occurred despite the
fact that more than 80% of the PCB mass in the dredging footprint was
estimated to have been removed by the dredging operation. Residuals
(generated and undisturbed) were measured at this site and an average
of about 16 in. of contaminated sediments remained after dredging
(range from 3 to 32 in.) (Connolly et al. 2006). Following dredging, the
dredged area was capped with an average of 1.5 ft of a sand and topsoil
mix. At this site, surficial sediment concentrations increased due to
dredging, although there was a large removal of PCB mass from the
river. The analysis illustrates that dredging can achieve substantial mass
removal but may not reduce surficial sediment concentrations. The data
were also analyzed for spatial correlation to identify and account for any
form of spatial variation in surficial PCB concentrations or mass (see Box
4-5).
127
BOX 4-5 Statistical Analysis of Surficial Sediment PCB Concentrations,
Data
PCB concentrations in sediment (mg/kg dry weight) from samples taken at
three times—before dredging, after dredging, and after capping—and from two areas—Main Channel and Northern Near Shore—were analyzed (EPA, unpublished
data, April 18, 2006). Longitude and latitude spatial coordinates for samples were also
available.
Methods
Linear regression models with an indicator of monitoring time (pre-dredging,
post-dredging, and post-capping) were used to statistically compare the sediment
PCB concentrations before and after dredging and after capping. The spatial sample
design (data coordinates) was not sufficiently consistent between times to consider a
repeated-measures-based approach. The Box-Cox transformation (Box and Cox 1964)
was parameterized in the regression-model likelihoods to allow possible transformations to be chosen optimally. Regression-model inference was based on maximum
likelihood. Analyses were stratified by geographic region and considered PCB concentrations. The nonparametric Kruskal-Wallis Rank Sum Test (Hollander and Wolfe
1973) was used to compare surficial PCB concentrations from the Northern Near
Shore because the sample sizes were small.
Results
Figure 4-13 displays the region-specific distribution and sample sizes for log10
surficial PCB concentrations. For sediment in the Main Channel, there was a significant increase in average PCB concentrations after dredging (p <0.01) and post-capping
concentrations were not significantly different from averaged pre-dredge concentrations. In the Northern Near Shore, the decline in average PCB concentrations between
pre-dredging and post-dredging was not statistically significant. After capping the
Northern Near Shore area, the average PCB concentration in surficial sediments was
significantly lower than before dredging (p <0.01). Regression analyses were on the
natural-logarithm-transformed data.
Remarks
The geographic layout of sample locations was not sufficient to allow statistical
models to identify and account for any form of spatial variation (either as a regression
128
BOX 4-5 Continued
trend or residual dependence) in PCB concentration. Residual spatial dependence is
known as potentially biasing tests of significance (Cressie 1991). As described in this
report, the entire main channel area was not able to be dredged. For the Main Channel analysis, pre-dredging, post-dredging, and post-capping data were compiled for
the dredged area only (referred to as “extended work zone 1”). Pre-dredging data
were compiled from three pre-dredging sampling periods for the dredged area
(termed “Phase II,” “January 2004,” and “Pre-ROPS,” generally top 3 in.). The postdredging (collected 10/22/2005; generally top 3 in.) and post-capping data (collected
11/28-29/2005; top 2 in.) from that area were also compiled. Northern Near Shore
samples were compiled from pre-dredging (9/9/2004, top 3 in.), post-dredging
(8/19/2005, top 3 in.), and post-capping (11/29/2005, generally top 2 in.) sampling efforts.
FIGURE 4-13 Distribution and sample sizes for Grasse River surficial sediment
PCB concentrations (mg/kg dry weight) stratified by geographic region—Main
Channel (MC) and Northern Near Shore (NNS)—and whether collected before
dredging (Pre), after dredging (Post), or after capping (Pcap). N = sample number. Further details are provided in Box 4-5.
129
General Motors Central Foundry, Massena, New York
In 1995, General Motors dredged an area of about 10 acres in the St.
Lawrence River near Massena, NY, that was contaminated with PCBs as
a result of the release of hydraulic fluids. The action removed over 13,000
cy of sediment and over 99% of the PCB mass in the sediment (EPA
2005a). However, it did not meet the cleanup level of 1 mg/kg in all locations, because there was residual contamination after dredging. As described in the remedial action completion report (BBL 1996), boulders
and debris were excavated mechanically, and sediment was removed
later with a horizontal auger dredge. The contaminated sediment was
underlain with dense glacial till that made it impossible to use overdredging to increase sediment-removal efficiency. In areas in which initial concentrations exceeded 500 mg/kg, 15-18 dredge passes were required to reduce sediment concentrations to below 500 mg/kg. In one
area that initially exceeded 500 mg/kg, eight additional attempts, including multiple dredge passes, were conducted to reduce sediment concentrations. Ultimately, the contractor concluded, with EPA concurrence,
that attainment of target cleanup levels in this quadrant was not possible
with dredging alone, and capping was instituted (BBL 1996). Without
capping, high residual PCB concentrations would have remained at the
sediment surface and limited the effectiveness of the remediation.
Manistique Harbor, Michigan
The presence of debris and bedrock limited the effectiveness of the
1995 to 2000 dredging operations to remove PCB contaminated sediment
at Manistique Harbor, MI (Nadeau 2006; Weston 2002). The primary
remediation goal of the project was the long-term protection of Lake
Michigan by removal of the potential PCB source in Manistique Harbor.
A secondary goal was reducing risks to people and wildlife that consume fish from the harbor (EPA 1994). As described in Weston (2002),
“Initially, the goal of the removal action was to remove sediments within
the dredge area with total PCB mass concentrations of more than 10 ppm
PCBs …The objectives of the removal action were further clarified and
restated …that the “objective of 95% removal of the total PCB mass from
within the AOC [Area of Concern] and an average concentration of not
130
more than 10 ppm throughout the sediment column within the AOC
shall be verified.” Remediation of PCB contaminated sediments began in
the fall of 1995 and was originally expected to continue through the fall
of 1997 and remove a total of 104,000 cy of sediment (MIDEQ 1996). Ultimately, the project required 6 seasons of dredging from 1995-2000 and
removed approximately 190,000 cy of contaminated sediment (EPA
2006a [Manistique River and Harbor Site, May 10, 2006]). The estimated
mass of PCBs removed by the end of the project was 82-97% of the initial
mass (Weston 2002). Information on pre-remediation surface sediment
concentrations varies. 16 Post-dredging average concentrations throughout the river and harbor (including dredged and non-dredged areas) reported for the top 1 ft were 9.0 mg/kg (sampled in 2000) and 7.3 mg/kg
(sampled in 2001) (Weston 2002). Nadeau (2006) summarized the pre-
EPA describes the results from two sampling events prior to dredging as
June 1993: “0-3 in.: Min: 0.15 ppm, Max: 124 ppm, Median: 3.4 ppm” and December 1993: “Min.: Below Detection, Max: 450 ppm” (EPA 2006a [Manistique River
and Harbor Site, May 10, 2006]). However, the area being referred to (for example, the whole harbor or just the dredged areas) is not specified. Post-dredging
evaluations of site conditions prepared for EPA also do not summarize predredging surface sediment concentrations in the dredged area (Weston 2002;
Weston 2005a). A summary by Malcolm Pirnie, Inc. and TAMS Consultants, Inc.
(2004) stated that predredging sediment sampling and characterization activities
indicated that the average PCB concentration in the top 3 in of sediment was 16.5
ppm. The 1996 Remedial Action Plan for the Manistique Harbor Area of Concern
(MIDEQ 1996) states “Sampling conducted in June and December 1993, April
1994 and May, June and July 1995, included most of the navigation channel,
along with other harbor and upstream locations…Sampling in the navigation
channel showed surface (0" to 3") concentrations of PCBs with a peak value of
120 ppm and an average of 16 ppm.” A feasibility study containing a review of
experiences at sediment dredging projects (BBL 2000) states “Pre-removal surficial sediment (0-3 in.) PCB concentrations in the Harbor ranged from non-detect
to 90 ppm (average of 14 ppm) using data collected during the Engineering
Evaluation/Cost Analysis (EE/CA) (BBL, 1994[a]). A consulting report for the
responsible party (BBL 1994b) that describes the predredging condition states
“The current surficial concentration for the 56 acre area used in the RA is 5.2 ppm
not the 8 ppm cited by [EPA]”. Several of the above averages are probably derived from the same data source (BBL 1994a), but the areas being considered in
that derivation likely differ (for example, the estimated dredged area vs. the 56
acre area of concern).
16
131
and post-dredging surface PCB concentrations as increasing slightly over
initial surface concentrations (concentrations increased from 5.2 to 7.3
mg/kg in the whole area17 and 15.1 to 18.8 mg/kg in the dredged area18).
Since completion of remediation, sediment from upstream has deposited
in the harbor area, burying sediments with elevated PCB concentrations.
A bathymetric analysis by EPA shows that in a five year period after
dredging (between fall 2000 and fall 2005), approximately 83,000 cy of
sediment deposited in the harbor from upstream sources, in some places
between 10-16 ft deep (EPA 2006a [Manistique River and Harbor Site,
May 10, 2006]). Surface sediment samples collected in 2004 (using a ponar dredge) had a mean concentration of 0.88 mg/kg PCBs in the area of
interest (Weston 2005a). This example illustrates both the difficulty of
eliminating residual sediment concentrations in the presence of debris
and bedrock and the inability to achieve long-term risk reduction because of the residual unless other processes, such as sedimentation, intervene to reduce surficial sediment concentrations.
Cumberland Bay, Lake Champlain, Plattsburgh, New York
Debris and a heterogeneous substrate caused dredging problems at
the PCB-contaminated Cumberland Bay, NY, site, where logs, wood
chips, and large rocks were encountered. Dredging began in July 1999
and ended in December 2000. After the initial dredging of the 34-acre site,
divers found many areas where PCB removal was incomplete, apparently because of the presence of debris. As described by the New York
State Department of Environmental Conservation (NYSDEC 2001), one
area originally thought to have been dredged to a hard bottom was
found by divers to be a hard crust that covered 4 ft of sludge containing
PCBs at 54 ppm. Hand-held hydraulic dredge lines were used by divers
to remove contaminated sediment from areas where dredging with only
the large hydraulic dredge was difficult. Another difficulty encountered
near a dock area was the bubbling up of gas during dredging that floated
sludge to the surface. On the completion of dredging, residual PCB con5.2 mg/kg: top 3 in., 1993 data, 56 acres; 7.3 mg/kg: top 12 in., 2001 data; 56
acres.
1815.1 mg/kg: top 3 in., 1993 data, 15 acre dredged area; 18.8 mg/kg: top 12 in.,
2001 data, 15 acre dredged area.
17
132
centrations averaged 6.8 mg/kg on the basis of analyses of 51 samples
taken from 42 cores. That is lower than pre-remediation PCB concentrations. A dock area had previously averaged 430 mg/kg, with maximum
of 13,000 mg/kg, and other dredged areas had averaged 33 mg/kg. However, the reduction in risk cannot be quantified, because no risk-based
numeric remediation goals were selected for the site.
Fox River, Wisconsin
Residuals were noted at three dredging projects on the Fox River.
At the 1998-1999 removal at Deposit N, sediment rested on a fractured
bedrock surface, so it was not possible for a dredge to cut into a clean
underlying layer. Sediment PCB concentrations after dredging averaged
14 mg/kg, similar to the average pre-dredging concentration of 16 mg/kg.
It is estimated that of the pre-project 142 lb of PCBs measured at Deposit
N, 111 lb was removed, and about 31 lb remained in the residual sediment on the completion of the project (Foth and Van Dyke 2000).
At the 1999 Fox River SMU 56/57 demonstration dredging project,
steep side slopes, debris, and underlying clay made it difficult to remove
contaminated residuals. Final cleanup dredging passes were performed
in four subareas before termination of the 1999 dredging, and postdredging surface concentrations in three of the four areas were less than
pre-dredging concentrations, although the 1-mg/kg target was not generally achieved even with cleanup passes. The overall post-dredging average surficial sediment PCB concentration at the end of 1999 dredging
was 73 mg/kg, compared with 4.4 mg/kg before dredging (Montgomery
Watson 2001). The surficial sediment concentration in the areas not subject to overdredging exhibited an average concentration of 116 mg/kg
and a median of 45 mg/kg, very close to the initial 53 mg/kg average
concentration in the deposit (Reible et al. 2003). Dredging of the remaining volume was completed by the responsible party as a removal action
in 2000 and achieved an average surficial sediment PCB concentration of
2.6 mg/kg; this was followed by backfilling with a minimum of 6 in. of
clean sand. 19 The outcome exceeded closure requirements for the reAccording to Fort James Corporation et al. (2001): “The vertical extent of the
dredging, as determined by the cleanup objectives, resulted in 28 of the subunits
being dredged to cleanup objectives, and two of the subunits dredged to develop
19
133
moval action, and surficial concentrations after the incorporation of
backfill were not measured (Fort James Corporation et al. 2001).
Early reports of the 2005 full-scale dredging of the upper reaches of
the Fox River have indicated problems in achieving cleanup targets (Fox
et al. 2006). The reports indicate that dredging did not remove all
sediment with PCB exceeding 1 mg/kg and that sand backfilling will be
necessary to meet the 0.25-mg/kg surface-area-weighted average
concentration end point. High-concentration deposits in thin soft sediment layers overlying stiff clay have made residual contamination
difficult to remove. A pilot test conducted in a portion of the dredged
area indicated strongly diminishing returns for redredging: doubling the
volume removed with the goal of removing all soft sediment above
native clay was not sufficient to meet the remedial goal of PCB at 1
mg/kg. Dredging results available at the time of the study were from
three subunits of OU 1 (Subunit A, C/D2S, and POG1). The preliminary
pre-remediation and post-remediation results—PCB mass removal and
surficial (upper 4 in.) PCB concentration—from verification sampling are
presented in Table 4-1.
More recently, a specialized dredge (a cutter-less head suction
dredge) developed by the remedial contractor was used during the 2006
dredging at the site. The dredge was designed for very thin deposits of
sediments over a clay or hard till bottom (including generated residuals).
Preliminary results for three dredge management units (about 2 acres
combined) show that concentrations well under the 1 mg/kg remedial
action level were attained even when initial concentrations ranged from
20 mg/kg to above 50 mg/kg (Green et al. 2007). The conditions (thin deposits over clay or hard till bottom) are considered to be among the most
difficult for attaining target cleanup levels.
Commencement Bay, Washington
The combination of the ability to overdredge into clean sediment
and the presence of sediment that has minimal debris or other obstacles
stable sideslopes for the dredge area. All 28 subunits met the cleanup objective of
10 ppm PCBs or less. Eleven of the subunits have PCB concentrations less than 1
ppm.”
134
TABLE 4-1 Summary of Pre-dredging and Post-dredging Verification Sampling
Results (2005) from Three Subunits of Operable Unit 1 in the Lower Fox River
Sub-area
Measure
Post-dredging Pre-dredging % reduction
A
PCB mass (kg) 205.5
26.6
87
Avg. surficial
PCB conc.
13.3
2.8
79
(ppm)
PCB mass (kg) 24.1
1.1
95
C/D2S
Avg. surficial
PCB conc.
7.6
1.0
87
(ppm)
PCB mass (kg) 36.2
1.3
96
POG1
Avg. surficial
PCB conc.
13.7
1.8
87
(ppm)
Note: PCB concentrations and mass for dredged area only, not entire subarea.
Source: Fox et al. 2006. Reprinted with permission; copyright 2006, Natural
Resource Technology, Inc. and CH2M Hill.
to dredging has led to more manageable residual concentrations during
the cleanup of several waterways in Commencement Bay, Tacoma, WA.
The 1993-1994 Sitcum Waterway cleanup in Commencement Bay was
combined with a redevelopment project by the Port of Tacoma designed
to create a capacity to handle deep-draft vessels in its facility. As a result
of the desire to increase navigable depth, the dredging plan included
removal of sediment to a bottom elevation that exceeded the depth of
contamination in open-water areas by at least 2 ft. The ability to overdredge facilitated removal of contamination in those areas and helped to
reduce the impact of contaminated residuals on final sediment quality.
Immediately after dredging in 1994, sediment quality objectives (SQOs)
had not been achieved in all areas. An additional 2 ft of sediment was
dredged from one of the areas, and sampling indicated that concentrations in the area were below the SQOs. In the other areas above the SQOs,
natural recovery was determined to be sufficient to meet the remedial
action objectives; these areas achieved SQOs in 2003. In 2004, EPA approved the Port of Tacoma's request to end further sediment monitoring
(EPA 2006a [Commencement Bay–Sitcum Waterway, April 26, 2006]).
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The 2004-2006 Head of Hylebos dredging project also indicated the
effectiveness of overdredging to reduce residual contamination. The
Hylebos Waterway was originally cut into a broad river delta consisting
of native sediments composed of clean and fairly compact silts and sands.
After the waterway was established, industry developed along the waterway, and this resulted in industrial-chemical discharges into it. No
river was feeding the waterway, so it slowly shoaled in with very finegrained sediment in the form of “soft black muck” over the natural or
native sediment. Characterization of subsurface sediment with core
samples showed that the contaminants from the industrial discharges
were restricted to the fine-grained surface sediment, whereas the immediately underlying native sediment was not contaminated (Dalton, Olmsted & Fuglevand, Inc. 2006). There was a clear visual difference between
the contaminated sediment and the underlying native sediment (compact silts and sands). During dredging, each bucket of material was examined visually by onboard inspectors to ensure that all fine-grained
sediment had been removed before moving on to the next area. In that
manner, overdredging into clean sediment could indicate that residual
contamination was minimal. The ability to differentiate clearly between
contaminated and uncontaminated sediment, dredging into the uncontaminated sediment, and the relative lack of debris combined to minimize the residuals.
Pre-dredging (dates unspecified) and post-dredging (August 2004January 2006) surficial-sediment (top 10 cm) total PCB concentrations
from the Head of Hylebos and a few samples identified as post-capping
(January 2006) samples were available for analysis (unpublished data;
Paul Fuglevand; Dalton, Olmsted & Fuglevand, Inc.; August 22, 2006).
Linear regression with an indicator of monitoring time (pre-dredging or
post-dredging) was used to statistically compare the sediment PCB concentrations before and after dredging. There was a significant decrease in
PCB concentrations in surficial sediment after dredging (p <0.01); the
pre-dredging mean was 685.9 μg/kg dry weight (n = 135), and the postdredging mean was 74.7 μg/kg dry weight (n = 400). Only six samples
were identified as post-capping samples, and they ranged in concentration from 36.4 to 847.0 μg/kg dry weight. This site remains one of the few
where cleanup levels were obtained by dredging alone (except in a few
areas). As stated by EPA, “dredging to expose clean native sediment was
successfully completed throughout the entire project area, with the ex-
136
ception of an under-dock cap (completed in 1998) and a shoreline subtidal cap completed early this year at a location of groundwater discharge with elevated arsenic concentrations. The sediment remediation
project successfully achieved the project SQOs with no residual sediments exceeding the SQOs (except at the two noted capping areas which
were not driven by generated residuals from dredging)” (EPA 2006a
[Commencement Bay - Head of Hylebos, May 17, 2006]). The comparatively low initial concentrations in the Hylebos Waterway (and many of
the Puget Sound sites) and the smaller magnitude of difference between
contaminated sediment concentrations and cleanup levels decrease the
potential effect of dredging residuals on achieving cleanup levels. However, generated residuals are derived at least partly from the contaminated material being removed so residuals management remains a critical issue at Pacific Northwest sediment sites.
Harbor Island, Duwamish River, Washington
Harbor Island is another site in the Puget Sound area whose stratigraphy is conducive to dredging (a clear separation between contaminated and native, largely uncontaminated sediment). Residual contamination after dredging may be more important than observed at the Head
of Hylebos or Sitcum waterways, because of extensive debris. The Lockheed Shipyard is on the eastern bank of the West Waterway in the lower
Duwamish River in the Harbor Island Superfund site. The site was the
location of a bridge-building company and then a ship-building and
maintenance facility. At the time of remediation, the site comprised a
failing bulkheaded shoreline; almost the entire nearshore area was covered by docks or marine railways, and an open-water area was immediately adjacent to the federal shipping channel. Extensive debris was present in the underpier and open-water area immediately adjacent to the
pier face. Surface debris included consolidated machine turnings and
other metal debris, cable, concrete blocks, and wood. Much of the site
was covered with thousands of deeply embedded creosote piles that
were slated for removal as part of the remedy. The remedy selected for
the site included dredging and capping of the nearshore area (not all
contaminated sediment would be removed) and removal of the contaminated sediment layer in the open-water area including debris that might
137
limit dredging effectiveness. Habitat restoration was a major component
of the remedy (EPA 2006a [Lockheed Shipyard Sediment OU, May 12,
2006]). Additional subsurface debris was encountered once dredging
began, including large concrete pier blocks, broken piles, and the original
willow cribbing that was used to contain dredged material during the
construction of the West Waterway and Harbor Island around 1900. Debris removal had an important effect on the duration, timing, and cost of
the project, and the remedy had to be implemented in two phases over
two seasons (because of restricted in-water work periods for the protection of endangered species in the Duwamish River). Phase 1 consisted of
pier and railway demolition, bulkhead replacement, and initial debris
removal (by dredging). Dredging to complete the debris removal and
achieve cleanup levels was conducted as phase 2 from November 2003 to
March 2004 and from October 2004 to November 2004 (EPA 2006a
[Lockheed Shipyard Sediment OU, May 12, 2006]). Because cleanup levels were not achieved after the first phase of dredging, a thin layer of
sediment (6-12 in.) was placed over the dredged area to stabilize the residuals until the next season of dredging, when it was removed as part of
the dredged inventory.
Summary
The available project data indicates that sediment resuspension and
the generation of residuals represent a nearly universal problem in connection with dredging of contaminated sites. Resuspension can be more
of a problem in the presence of debris or other site conditions that interfere with normal dredging operations. In addition, readily desorbable
contaminants and fluid contaminants, such as nonaqueous-phase liquids,
are unlikely to be effectively captured by the dredge or by common operational controls on resuspension. Nor will such contaminants be adequately characterized by measuring the suspended solids, such as TSS or
turbidity.
Low sediment bulk density and the presence of debris and hardpan
or bedrock all tend to increase resuspension and residuals. Available
data indicate that dredging is most likely to be successful when dredges
penetrate into clean sediment layers reducing the amount of generated
residuals. At sites where structures, debris, hardpan, or bedrock limit
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dredging effectiveness, the desired cleanup levels, if based on the attainment of specified chemical concentrations, are unlikely to be met by
dredging alone. The inability to attain cleanup levels would presumably
translate into an inability to meet both short-term and long-term remedial goals and objectives.
Resuspension appears to result in at least short-term negative impacts on water quality and organisms. Residuals may give rise to longer
term negative impacts, but at most sites, there has been insufficient
monitoring to evaluate the long-term impact of residuals or capping has
been used to manage residuals.
MANAGEMENT OF DESIGN AND IMPLEMENTATION TO
MAXIMIZE DREDGING EFFECTIVENESS
Although the factors affecting dredging effectiveness outlined in
the preceding section are operative at all sites, their influence can be
minimized, although not eliminated, by active management of the
dredging process, that is, managing design and implementation to
maximize effectiveness. Through experience gained at dredging sites, a
number of actions have been identified that can help to maximize the
effectiveness of dredging in particular situations. However, that experience also suggests that successfully overcoming the limitations of dredging requires both site conditions conducive to dredging and the implementation of some or all of those actions. Sites that exhibit extensive
debris, hardpan or bedrock immediately below contamination, or other
factors that limit the ability to control residuals or resuspension will continue to be problematic for dredging even if all the actions discussed below are implemented.
Ensure Adequate Site Characterization
Central to the successful implementation of any remedial action is
site characterization sufficient to define a conceptual site model. A comprehensive conceptual site model should define the contaminants of concern at a site, the spatial distribution of contamination, the processes that
describe the change in contamination over time, the human and ecologic
exposure routes, and the significance of exposure and risk. Only when
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these aspects of the model are developed can a remedial effort be designed to respond to risk appropriately and achieve remedial goals.
Adequate site characterization can identify potential sources of contamination and provide the data necessary to design an effective remedial
program.
At the Reynolds Metals Superfund site on the St. Lawrence River
near Massena, NY, pre-dredging site characterization was not adequate
to delineate the distribution of the chemicals of concern at the site.
Dredge design was based on the assumption that the PCBs were collocated with the other chemicals of concern, PAHs, and total dibenzofurans. However, post-dredging sampling indicated that this was not the
case (EPA 2006c). Following dredging, which included redredging several of the areas, it was determined that PCBs were not collocated with
PAHs and that about one-third of the 22 acre dredged area contained
PAH concentrations above the cleanup level (EPA 2006c).20 Future remedial activities at this site are currently being decided.
The Head of the Hylebos Waterway was adequately characterized
prior to dredging. Historical surface and core samples were used in conjunction with planned studies to determine the horizontal and vertical
distribution of contaminants. As described in the remedial action construction report (Dalton, Olmsted & Fuglevand, Inc. 2006), the historical
U.S. Army Corps of Engineers (USACE) post-dredging surveys were
used to map the interface between the soft black muck and the native
bed sediments and to refine the dredge plan. However, core samples
were used to confirm the interfaces, and care was taken not to composite
core samples across the muck-native interface. Over 100 cores and over
500 surface samples were used to delineate the area and depth for remediation in this approximately 45 acre site. The coring studies were also
used to establish that the recent and native sediments were physically,
visually, and chemically different from each other and that chemical exDredging to remove PCBs was more successful although one area required
post-dredging capping. As described by EPA (2006c): “Despite extensive dredging of the St. Lawrence River, the cleanup goals of 1 mg/kg PCBs…were not
achievable in all areas. As a result, a 0.75-acre, 15 cell area, containing a range of
PCB concentrations from 11.1 mg/kg PCBs to 120.457 mg/kg, was capped with
the first layer of a three-layer cap to achieve the cleanup goal. The remaining
exposed sediments average 0.8 mg/kg PCBs within the remaining 255 cells (21
acres), which is below the cleanup goal.”
20
140
ceedances of the SQOs were only found in the recent sediments (EPA
2006a [Head of Hylebos Waterway, Commencement Bay, May 17, 2006]).
During implementation, the designers viewed the deepest historical
dredging as a general guide but relied on observations during dredging
to establish successful removal of the impacted sediment.
The design of a dredging plan requires interpolation of depth-ofcontamination data from sediment core samples. The upstream portions
of the Fox River (OU 1), where dredging is currently being conducted, is
a challenging site in that regard because it consists of multiple discrete
contaminated sediment deposits arising out of local differences in flow
regime and relationship to contaminant sources. The method applied by
the remedial design team in OU 1 was to develop deterministic interpolations of sediment PCB concentrations for each deposit and then to connect the interpolations at deposit boundaries (CH2M Hill 2005). When
that method is applied, the result is a surface of predicted depth of contamination, which can be expected to be most accurate in the neighborhoods of samples used in the interpolation and most uncertain in unsampled locations and at boundaries of deposits.21 The spatial density of
cores is a key component of adequately characterizing sediment distribution (particularly cores that penetrate through the entire deposit). However, there is no single optimum spacing between core samples because
the necessary density depends on the heterogeneity of the site deposit
and is site specific.
Accurate characterization is particularly challenging in areas where
contaminated sediments are underlain by uneven sub-bottom (for example, furrows, gulleys, or depressions). In these areas, deposition of contaminated sediment over time will often fill in low spots to create a relatively flat sediment-water interface, but with marked differences in the
underlying depth of contamination (for example, see description in Box
4-5). At the afore-mentioned Cumberland Bay site, variations in the uncontaminated sub-bottom characteristics and topography proved difficult to characterize prior to dredging. According to the NYSDEC, “site
characterization and pre-design studies included bathymetric surveys
Alternative interpolation methods exist, such as the geostatistical technique
of kriging, providing the probability of contamination at various depths at every
location, and interpolating continuously across deposits while allowing for deposit-specific effects (Cressie 1991; Goovaerts 1997; Diggle and Ribeiro 2007).
21
141
and sludge probing to define the top of the sludge bed, its thickness, and
horizontal extent. In addition to sludge coring, divers confirmed the
outer extent of the sludge bed in areas where it was too thin to measure
by coring.” However, following dredging in 1999, it was determined that
in one area “originally believed to have been dredged to a hard bottom
since the sampling device encountered refusal” the sediment “consisted
of a hard crust underlain by up to four feet of [contaminated] sludge.” In
another area, PCB-contaminated sludge was found in 1-6 ft deep depressions scattered along the bottom of the lake following dredging. Further
dredging targeted both of these contaminated areas (NYSDEC 2001).
In the previously described 2005 pilot study in the Grasse River, it
also proved difficult to accurately define the thickness of contaminated
sediments using available sampling techniques and protocols. Prior to
dredging, multibeam bathymetry, sediment probing, sediment coring,
and acoustic sub-bottom profiling were used to characterize the site.
Depth to hard bottom was estimated using sediment probing on a 25-ft
by 25-ft grid with PCBs expected to be present in sediments above the
hard bottom (Connolly et al. 2007). Following dredging, vibracore sampling indicated that in some areas the estimated thickness of contaminated sediments was wrong and significant contaminated sediment remained (see Figure 4-14). These results indicated that at this site
“sediment probing is not a reliable indicator of the depth of sediment
and manual pushcore sample collection is not a reliable indicator of the
full depth of PCBs contaminated material and below the deepest contaminated layer” (EPA 2006a [Grasse River Site, April 18, 2006]). Acoustic subsurface sampling at this site was also not successful for detailing
sub-bottom characteristics.22
A similar situation existed at the Manistique River and Harbor site.
During site characterization, the samples taken before the dredging were
According to EPA: “Sub-bottom profiling attempted in conjunction with
multibeam bathymetry survey on October 22, 2005. Used dual frequency Odom
depth sounder to obtain sounding information along several transects situated
parallel to direction of river flow. 200 kHz signal reflects off sediment surface
and 24 kHz signal penetrates into sediments and bounces off sediment reflectors.
No penetration beyond reflections from sediment surface achieved” (EPA 2006a
[Grasse River Site, April 18, 2006]).
22
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FIGURE 4-14 Upstream to downstream transects of the Grasse River remediation
area showing: elevations of the dredging cut line (dotted lines) based on predredging (Spring 2005) depth of contamination estimates; post-dredging (Fall
2005) extent of contamination (diamonds) determined by vibracore sampling;
and the post-dredging (Fall 2005) bottom elevation determined by bathymetry
(solid lines). Source: Connolly et al. 2007. Reprinted with permission from the
authors; copyright 2006, Quantitative Environmental Analysis.
thought to be taken to bedrock, but were not. This is apparently because
wood debris under the sediments was thought to be the bedrock harbor
bottom (EPA 2006a [Manistique River and Harbor Site, May 10, 2006]).
At this a site, a subbottom profiling device was used to estimate sediment thickness. However, the wood pulp and debris in the sediments
contained large amounts of gases that rendered the subbottom profiling
device useless. As a result, EPA indicated that the dredging depth could
not be predetermined in each area (EPA 2006a [Manistique River and
Harbor Site, May 10, 2006]).
The influence that incomplete characterization has on the success of
dredging points to the importance of accurate site characterization.
However, characterization activities are resource intensive and can consume time and funds otherwise available for remedial activities. As a
result, decisions on whether to proceed with further characterization
should seek to ascertain whether additional characterization will benefit
remedial effectiveness and the point at which the additional efforts provide diminishing returns. These considerations will be site-specific. Obviously, areas with complex and heterogenous sub-bottom (such as the
143
Grasse River) will benefit from greater characterization than those with
less heterogeneity. The variety of subsurface characteristics at these sites
also indicates that the most useful characterization technologies will be
site specific. Sediment sampling coupled with an understanding of the
fluvial and geologic nature of an area will shed light on the attributes of
the sediment deposit, but not all site conditions can be completely understood prior to beginning work. As a result, verification samples and
progress cores taken during dredging are useful for indicating whether
operations are succeeding or modifications need to be made (see Chapter
5 for further discussion).
Defining ongoing sources of contaminants to the waterway
through site characterization is of critical importance for determining
appropriate cleanup responses and for eliminating recontamination of
remediated areas. This issue is addressed further in the next section.
Implement Source Control
As pointed out by the National Research Council Committee on
Remediation of PCB-Contaminated Sediments, “the identification and
adequate control of sources of PCB releases should be an essential early
step in the site risk management” (NRC 2001). If contaminant sources are
not controlled, dredging cannot be effective in managing risk. In the
Hylebos Waterway, the combined efforts of EPA and the Washington
State Department of Ecology to achieve source control before dredging
contributed to the ability to implement successful remedies. Dredging of
the Hylebos Waterway was also at least partially successful because of
the efforts to directly address inaccessible areas that could not be
dredged. Much of the shoreline of the waterway is modified with overwater structures, such as docks, piers, and wharves. The remedy selected
for the head of the waterway included dredging accessible areas, excavation or capping in isolated intertidal and under-pier areas, and natural
recovery. Without control of the contaminated sediments under piers
and along the shoreline, the project would likely not have been considered successful. Dredging beneath the Arkema dock during 2005 used a
long-reach excavator to remove most of the impacted sediment followed
by a diver-deployed hydraulic dredge to remove the loose residual material that accumulated during mechanical dredging. The 2005 dredging
144
activities achieved the SQO cleanup objectives beneath both the Ace
Tank and Arkema docks. Some areas along the shore were inaccessible to
dredging and were either capped or determined to be suitable for monitored natural recovery.
In contrast, both potentially important source areas and exposed
sediment contributing to risk were missed, or not appropriately evaluated, in the characterization of the Lauritzen Channel at the United
Heckathorn site. Dredging did not achieve remedial action objectives at
this site despite achieving cleanup goals immediately after dredging for
DDT, the contaminant of concern—apparently because of a failure to
address contaminant sources or mass contributing to exposure and risk
at the site. Confirmation sampling after dredging seemed to confirm the
cleanup goal of 0.59 mg/kg in the Lauritzen Channel of the site, but an
investigation a year later, in 1998, found DDT concentrations as high as
30.1 mg/kg. Year 1 biomonitoring showed that pesticide concentrations
in the tissues of mussels exposed at the site were higher than those observed before remediation; these values decreased slightly in later years.
Anderson et al. (2000) noted remaining toxicity in amphipods after
dredging. In 2002, a buried outfall visible during low tide was identified
as a persistent source of DDT in the channel; it was plugged by EPA in
2003. Investigations in 2002 and 2003 found sediment concentrations
greater than 1,000 mg/kg in the vicinity of a dock on the eastern side of
the channel. Remedial objectives were not met in the Lauritzen Channel
for a number of reasons, including the decision not to dredge side slopes
completely or to remove material from under piers. Some of those areas
were intended to be capped with sand, but because of the steepness of
slopes or inaccessibility, this was not done (Chemical Waste Management 1997).
A recent analysis of recontamination of completed sediment remedies based on publicly available reports, such as 5 Year Reviews, indicated that 20 areas where dredging or capping remedies had been completed were recontaminated from outside sources, primarily by
combined sewer outfalls (CSOs), unremediated upland areas, and adjacent and upstream unremediated areas (Nadeau and Skaggs 2007). The
potential for recontamination at a site underscores the importance of
identifying and controlling sources before undertaking a sediment remedy.
145
Monitor Appropriate Indicators of Effectiveness
Adequate site characterization should provide an understanding of
the key sources of exposure and risk at a site and of how to intervene
effectively to control risk. It should allow the definition of appropriate
remedial action objectives and of cleanup levels that will lead to their
achievement. It should also therefore identify appropriate indicators of
successful implementation and, ultimately, effectiveness. A baseline of
appropriate indicators of effectiveness must be established before dredging to make possible a comparison with post-dredging data. Monitoring
should continue until effectiveness can be evaluated.
At the Outboard Marine Corporation—Waukegan Harbor site, success of the dredging remedy was monitored by using, among other indicators, fish-tissue data. The site is in Lake County, IL, 50 miles north of
Chicago, and consists of industrial, commercial, municipal, and open or
vacant lands at the mouth of the Waukegan River and North Ditch
drainage basins. It is estimated that 300,000 lb of PCBs (Aroclors 1242
and 1248) were released into the harbor and that sediment concentrations were up to 25,000 ppm (EPA 2000a). The harbor was dredged in
1992 and contaminated sediments were placed in an abandoned boat slip.
Only areas exceeding 50 ppm were remediated; as a result, some areas in
the harbor are expected to have relatively high residual concentrations,
and further remediation is being considered (EPA 2002; EPA 2007b).
Fish-tissue data from Waukegan Harbor were analyzed to evaluate
the hypothesis that PCB contamination has decreased (see Box 4-6). To
evaluate dredging effectiveness accurately on the basis of preremediation and post-remediation fish body burdens, data sufficient to
estimate both a pre-dredging time trend and a post-dredging time trend
are needed from representative samples of fish collected from exposure
areas that are the subject of cleanup levels and remedial action objectives.
Effectiveness can then be shown if the post-dredging trend is lower than
would be expected from simple extrapolation of the pre-dredging natural-recovery trend. For Waukegan Harbor, sampling was conducted at
only two times before dredging. That pre-dredging time trend was inadequate for comparing to the post-dredging time trend.
A pre-dredging sediment toxicity test found relationships between
toxicity and sediment-contaminant concentrations, with toxicity ranging
146
BOX 4-6 Statistical Analysis of PCB Concentrations in Fish,
Waukegan Harbor, Illinois
Data
Analyzed data (T. Hornshaw, Illinois EPA, written commun., August 3, 2006)
included percent lipid and total PCB concentrations in carp tissue (fillets) that were
caught before dredging (1981 and 1983) and after dredging (1996-2001 and 2005).
Dredging activity was conducted during 1991-1992.
Methods
Temporal analysis of fish trends was based on linear-regression models of PCB
concentrations in fish samples with monitoring year and percent lipid content as independent variables. The Box-Cox transformation (Box and Cox 1964) was parameterized in the regression-model likelihoods to allow possible transformations to be
chosen optimally. Nonlinear trends in time were considered (Stow et al. 1995), and
their results led to interpretations that were qualitatively similar. The two-sample
Wilcoxon Rank Sum Test (Hollander and Wolfe 1973) was used to compare percent
lipid-normalized PCB body burdens before and after dredging.
Results
Figure 4-15 displays the log base 10 lipid-adjusted total PCB in carp fillets for
the available monitoring years. There was no significant difference between mean
pre-dredging (1981 and 1983) lipid-normalized PCB and mean post-dredging (19962001 and 2005) lipid-normalized PCB (p = 0.34). Despite the scatter of the data points,
there is a statistically significant 12% decline in lipid-adjusted PCB per year for 19962005 (p = 0.03).
Remarks
The temporal trend shown in Figure 4-15 was established only on the postdredging data. Comparisons with the pre-dredging monitored fish (total sample size,
7) were insufficient to establish any conclusions on dredging effectiveness. Improved
longitudinal monitoring (before and after dredging) could provide data sufficient to
establish a time trend that could be used to associate changes with dredging. Monitoring during dredging would provide insight into the effects of sediment release and
resuspension and inform statistical models as to when to postulate post-dredging
effects.
The comparison of pre- and post-remediation data is further complicated by
likely improvements in analytical procedures between 1985 and 2005. Other issues,
for example, the failure to segregate fish by age or size, are also likely to affect this
analysis.
147
FIGURE 4-15 Lipid-normalized PCB concentrations in carp in Waukegan Harbor,
IL. Dredging was completed in 1992. See Box 4-6 for details of the regression
analysis.
from 0 to 100% survival for Ceriodaphnia dubia and Daphnia magna in
whole-sediment laboratory assays. Hyalella azteca survival was much better in 48-hr exposures with survivals of 73.3 to 100 % (Burton et al. 1989).
A post-dredging study by Kemble et al. in 2000 suggested that the PCB
concentrations were lower (less than 10 mg/kg) and that sediments were
generally not lethal to amphipods, but there were sublethal effects. The
Kemble et al. (2000) and EPA (1999) studies suggested that because postdredging PAH concentrations in sediments exceeded sediment quality
guidelines based on probable effect concentrations, they may be contributing to the observed toxicity. The presence of PAH-related toxicity at a
site with cleanup levels based on PCBs points to the need to focus on the
full range of chemicals that may be causing toxicity at a site.
At the PCB-contaminated Cumberland Bay NPL, NY, site where 34
acres was dredged, quantitative remedial goals were not set. Rather, according to the site’s ROD (NYSDEC 1997), “The goals selected for this
site are: mitigate the immediate threat to the environment posed by the
PCB contaminated sludge bed; rapidly and significantly reduce human
health and environmental risks; [and] prevent further environmental
148
degradation resulting from this known source of PCB contamination.”
Without the guidance of risk-based quantitative criteria, it is difficult to
judge whether the remedy achieved the goals that were set. As mentioned previously, PCBs averaging 6.8 mg/kg were still present on completion of dredging.
At the Fox River, a baseline monitoring plan is under development
and is expected to establish the framework for long-term monitoring.
However, the plan is being developed after remedial dredging has begun. A wealth of data is available from the remedial investigation and
from scientific studies of the Lower Fox River and Green Bay that have
been conducted over the decades. Nevertheless, the initiation of the
dredging remedy before the establishment of a baseline—which ideally
would use media, methods, and locations consistent with the long-term
monitoring to follow—will probably complicate the evaluation of remedy effectiveness. In particular, it will be especially difficult to infer the
initial 5-year effects of dredging on contaminant exposures by comparing years 0 and 5, because the apparent baseline will also be affected by
dredging rather than reflect true pre-dredging conditions. Whether
dredging causes an immediate drop in exposures by removing contamination or causes an increase in exposures to contaminated residuals or
water-column releases, the collection of baseline data after the beginning
of dredging will confound the comparison of long-term monitoring data
with the true baseline. That will be true not only of the upstream portion
of the Fox River, where dredging is being implemented, but also of
downstream locations because upstream PCB releases to the water column can affect downstream conditions.
A common problem in monitoring for effectiveness is focusing on
the meeting of cleanup levels, especially if operationally defined, and not
on the long-term remedial action objectives. For example, cleanup for the
Christina River was based on sediment removal. Sediment contaminant
concentrations or biologic responses were not reported immediately following dredging or after backfilling (URS 1999). 5 years after dredging,
data on the status of the benthic community was collected (EPA 2005b);
however, this information is unable to indicate whether the remediation
was effective.23 Although the cleanup requirements based on dredging to
In the second 5-year review for the site (EPA 2005b), EPA states that “the
benthic community is dominated by pollution tolerant species that can be found
23
149
a specified elevation may have been met, attainment of short-term and
long-term risk-reduction goals has not been demonstrated. Similarly,
cleanup requirements were met at the Naval Shipyard in Newport, RI,
by removal of sediment to bedrock in many locations (TetraTech 2004).
However, dredging to the specified depth (bedrock) is not an appropriate indicator of risk reduction. There was no verification sampling in
these locations prior to capping (backfilling), so residual contamination
may remain. As a result, contaminants may also be available for transport through the sand cap to surface sediments. The observation of continued toxicity to sea urchins during long-term monitoring suggests that
risk-reduction goals were not achieved (TetraTech NUS 2006).
Evidence from other sites shows how meeting of cleanup levels,
even if based on post-dredging concentrations, might not achieve desired risk-reduction goals. The 1995 remediation at GM Massena, NY,
was designed to ultimately reduce exposure of fish and other wildlife to
PCBs in the sediment. As indicated previously, cleanup levels were
achieved only after capping of a portion of the site where reduction in
residual concentrations below the cleanup level was not achievable
solely through dredging. Examination of monitoring designed to evaluate performance relative to long-term remedial action objectives, however, has not shown expected reductions in fish concentrations. As
shown in Table 4-2, spottail shiner, a fish with a limited foraging range,
showed no obvious increasing or decreasing trends in PCB concentrations at the site even 5 years after the end of remediation (EPA 2005a).
Use Cleanup Levels Appropriately in Determining
Remedy Effectiveness
When comparing post-remediation concentration data to cleanup
levels, risk managers sometimes treat the cleanup levels as concentrations that should never be exceeded. However, this approach is not necin naturally stressed freshwater systems. One of the lines of evidence used to
determine that the river needed cleanup was the abundance of pollution tolerant
species. Since these areas have been dredged and backfilled with clean sediments, the prevalence of pollution tolerant species would not be attributable to
site related contaminant toxicity (there is still zinc in the surface water from other
sources).”
150
essarily appropriate or consistent with the evaluation of human and ecologic exposure conducted in the baseline risk assessments and, more importantly, with the derivation of cleanup levels. EPA guidance (EPA
1989b) recommends use of arithmetic mean concentrations within each
exposure area to quantify exposures to chemicals of concern over time.
While concentrations can vary significantly within an exposure area, the
arithmetic mean is the appropriate statistic based on the assumption that
a receptor integrates its exposure by moving about within the exposure
area. Surface area weighted average concentrations, for example, at the
Fox River, have also been used as appropriate indicators of exposure to
surficial sediments (WI DNR/EPA 2002). The Marathon Battery site illustrates the importance of comparing cleanup levels to the appropriate statistic. At this site, occasional sediment samples had concentrations that
were indistinguishable from the pre-remediation concentration distribution, yet these samples apparently are not reflected in the cadmium body
burden in the benthic community, which has decreased.
Because a sampling program provides an imperfect measure of the
arithmetic mean, EPA recommends use of the 95% upper confidence
limit (UCL) of the mean. Therefore, cleanup levels ideally should be
compared to the 95% UCL for monitoring data representative of the
exposure area of concern that incited establishment of the cleanup level.
EPA used 95% UCLs calculated from surface-sediment samples collected
after completion of dredging to determine whether cleanup levels had
been achieved at the Sitcum Waterway in Puget Sound (EPA 2006a
[Commencement Bay–Sitcum Waterway, April 26, 2006]). For the Sitcum
TABLE 4-2 Spottail Shiner PCB Concentrations After Remediation of the GM
Massena, NY, Site in 1995
Total PCBs-Whole
LipidBody Concentration
Normalized PCB
Number of
(mg/kg-lipid)
Date
Samples
Lipids (%) (mg/kg)
10/97
10/98
10/99
10/00
10/01
7
7
7
7
7
Source: EPA 2005a.
5.58
4.24
9.22
11.4
5.00
1.20
3.59
2.43
1.5
3.7
22
79
27
13
75
151
Waterway site, EPA reported that the “general approach [was to] redredge if [concentrations] exceed SQOs, but [EPA] also looked at
surrounding data, historical data, and 95th percentile UCL of the mean
sediment concentration for a chemical in a subarea to recommend
whether additional sampling should be done, re-dredging, natural
recovery, and/or nothing.” In cases in which monitoring data are biased
such that they are not representative of exposure areas of concern, one
can perform spatial weighting of the data before calculating 95% UCLs.
For example, EPA used an interpolation method called inverse distance
weighting to spatially weight floodplain soil data affected by
contaminated sediment in the Housatonic River before calculating 95%
UCLs for use in the human health risk assessment (Weston 2005b,
Attachments 3 and 4).
Consider Using Pilot Tests
As described above, adverse site conditions may significantly limit
the ability of dredging to achieve cleanup levels and remedial action objectives. Pilot testing can assist in identifying and characterizing potential
limitations to dredging effectiveness, planning and responding to unexpected factors that may arise, and in defining the degree of effectiveness
that might be obtainable through dredging. At the Lockheed Shipyard in
Puget Sound, delays, additional costs, and limitations of dredging effectiveness were encountered owing to the unexpected quantity of debris.
In retrospect, the construction manager for the remediation project stated
that a pilot dredging program would have been able to inform him of the
extent of the debris and the implications for dredging production rates
and rehandling issues (G. Gunderson, TRC Solutions, personal commun.,
July 7, 2006). Similarly, the dredging remedy at Manistique Harbor, MI,
was much more expensive, took twice as long, and required the dredging of nearly twice the expected sediment volume because of the unexpected problems associated with the presence of debris and thin sediment layers over bedrock. A pilot dredging program would have been
advantageous in efficiently characterizing the scale and costs of the
dredging remedy before full-scale implementation.
Pilot testing has been successfully used at a number of sites in identifying potential limitations to remedial effectiveness and allowing the
152
development of appropriate responses. For example, pilot testing has
been used at the Grasse River in Massena, NY, in which the effectiveness
of different dredges and different remedial technologies was explored.
The previously described demonstration dredging project at Lavaca Bay,
TX, was designed to evaluate the ability of dredging to effectively and
economically address mercury-contaminated bed sediment at the outfall
of a former chloro-alkali manufacturing site (Alcoa 2000). Results of the
study indicated that hydraulic dredging could be readily implemented at
this site, offsite transport of mercury on tidal flows moving through and
around the curtained-off dredging unit were minimal, a large mass of
mercury (2,300 lbs) was extracted from the hot spot, and increased mercury concentrations in oysters above the historical observed background
in the wider bay did not occur. Residual surface sediment (generally, 0-5
cm) mercury concentrations were reduced in areas with high surface
concentrations and lower subsurface concentration, while areas with
highly contaminated buried sediments and low pre-dredging surface
sediment contaminant concentrations typically showed increased surface
sediment concentrations post-dredging (Alcoa 2000). The pilot study was
judged to be a successful undertaking in that the data collected were key
in the evaluation of the role of dredging in the remedial activities to be
undertaken at this site.
Several dredging demonstration projects were conducted during
the remedial investigation and feasibility study at the Fox River. The
previously-described demonstration projects conducted in Sediment
Management Units 56 and 57 (SMU 56/57) in 1999 and 2000 provided
valuable information on dredging and dewatering productivity and operations. The 1999 demonstration removed 31,000 cy which was much
less than the 80,000 cy objective. During that project, hydraulic dredging
equipment was upgraded three times in an effort to increase the solids
content of the dredged slurry. Dewatering of solids proved to be a limiting constraint on production rate and required installation of additional
filter presses. The average production rate was 294 cy/day, compared
with a desired rate of 900 cy/day (Montgomery Watson 2001). Further
adjustments in dredging and dewatering equipment, beyond those made
by the 1999 project team, were needed in 2000 to remove the remaining
50,000 cy of targeted sediments. The resulting average production rates
exceeded a project target of 833 cy/day and reached a peak production
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rate of 1,599 cy/day on a single day near the completion of the removal
action (Fort James Corporation et al. 2001).
Pilot studies also assisted in the success of the Head of Hylebos
dredging project. Two pilot studies were used to help to define the scope
of such problems as debris, provide large samples for additional testing,
and help with selection of equipment and development of the dredging
operation plan. One study concentrated on how to remove the “soft
black muck drainage water” from the mechanically dredged material at
an upland storage facility. About 5 gal of water per cubic yard of sediment, or about 40 lb per 2,200 lb, was associated as free water. The second study focused on rail transport and placement into the offsite landfill and helped to refine the rail-transport program. During the pilot
study, the contractor observed the dissociation of physical integrity
(strength) of the soft fine-grained sediment after handling, both in the
barge and on the bottom of the waterway. The loss of strength was seen
as a contributor to formation of a flowable residual layer (fluidized soft
mud) on the bottom during dredging that needed to be captured (P.
Fuglevand; Dalton, Olmsted & Fuglevand, Inc.; personal commun., July
7, 2006).
In some cases, pilot testing is not required, because of the scale of
the dredging project or because of other conditions. For example, the
relatively small scale of the U.S. Naval Shipyard site in Newport, RI, and
the ability to dredge much of the contaminated material from land combined to make the implementation of dredging favorable. Of 30,000 cy
dredged, the vast majority was from the nearshore area (TetraTech NUS
2006). The nearshore materials were removed by a long reach excavator,
which was operated on a bay haul road constructed for the project, and
were loaded directly into offroad dump trucks (Tetra Tech 2004). Dredging of the remaining elevated offshore material was performed from a
barge with a crane equipped with a clamshell bucket and loaded onto an
adjacent haul barge. The excavator was much faster and less expensive
than the barge-mounted crane, and direct loading into haul trucks minimized handling of material (EPA 2006a [Newport Naval, May 17, 2006]).
In most contaminated sediment megasites, however, the scale and
complexity of the sites suggest that pilot studies are appropriate and will
assist in reducing limitations of dredging effectiveness. Adaptive man-
154
agement 24 and even pilot testing during implementation may still be
necessary to respond to unforeseen problems during implementation.
Pilot testing alone will not ensure success; to maximize the project’s usefulness, the scope and objectives need to be clearly communicated and
monitoring needs to be capable of establishing whether objectives were
achieved and the factors that influenced the project’s performance.
Implement Best Management Practices
Although it is not a guarantee, the adoption of best management
practices (BMPs) will help to ensure appropriate implementation of a
remedial project. Best management practices are defined on an activityspecific basis and will depend upon the type of dredging and transport
equipment used, the environment in which the dredging takes place, and
the process “train” or sequencing of the remedial activities. There are no
standardized BMPs for environmental dredging, although “lessons
learned” from environmental dredging projects to date suggests that
there are BMPs that will likely be applicable to many dredging projects.
These BMPs are primarily designed to minimize the loss or transport of
contaminated sediment or debris from the dredging footprint and minimizing the generation and runoff of leachate from dredged material to
the receiving water during transport or rehandling of dredged sediment.
BMPs that may be useful for minimizing loss or off-site transport of
sediment and debris include (this list is not considered comprehensive):
• Use of silt curtains to reduce the transport of suspended solids.
• Use of floating and/or absorbent booms to capture floating debris or oil sheens.
• Reduction of the impact speed of the dredge bucket with the bottom and/or reduction of the rate of ascent of a filled bucket; reduction of
the swing rate of cutter-head dredge.
• Prevention of overfilling buckets through accurate and controlled placement of bucket.
In general, adaptive management is the testing of hypotheses and conclusions and re-evaluation of site assumptions and decisions as new information is
gathered. See Chapter 6 for further detail.
24
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• Use of environmental or sealed buckets, where sediment characteristics will allow.
• Protection of the overwater swing path of a filled bucket (by
placing an empty barge or apron to catch lost material).
• Eliminating bottom stockpiling of dredged material or sweeping
with the dredge bucket/head.
BMPs that may be useful for controlling production or runoff of
leachate include:
• Maximization of the “bite” of a dredge bucket (that is, avoiding
thin lifts).
• Allowance for draining a sediment-filled bucket before breaking
the water’s surface.
• Use of filtration cloth, hay bales, curbing, or other physical baffles (similar to stormwater BMPs) to control runoff from barges or rehandling areas.
Additional BMPs that may be used to minimize environmental impacts of dredging include (but are not limited to):
• Scheduling dredging during periods when sensitive species or
populations are not present at the site.
• Daily construction oversight and progress surveys.
• Water quality monitoring during dredging activities.
BMPs for control or prevention of resuspension or loss of contaminated material in the waterway were implemented during the remediation of Todd and Lockheed Shipyards. BMPs were specified for overwater demolition, pile removal, dredging, barge dewatering, vessel
management, sediment offloading, capping and fill placement, and
overwater construction. During demolition, pile removal, and overwater
construction, an absorbent boom with 4- to 6-ft silt curtains was deployed to contain floating debris or sheen caused by the removal of creosoted piles. Entrainment of water during dredging was minimized by
taking complete “bites” with the dredge bucket whenever possible. Each
full bucket was held just at the water’s surface to allow water to drain
before the bucket was swung to the barge. Dredged sediment was pas-
156
sively dewatered on an onsite flat-deck barge through straw bales and
filter fabric before being discharged to the waterway. During offloading,
the clamshell bucket was prevented from swinging over open water by
placement of a spill-collection platform under its path. Asphalt curbing
surrounded the transloading area to prevent sediment, sediment drainage water, and contact stormwater from migrating offsite. Water collected from the transloading area was not allowed to enter the waterway
but was collected and treated on site by a process of settling, multimedia
filtration, and carbon filtration. Treated water was discharged to the
sanitary sewer.
Implementation of BMPs for control of produced water was also
important in the success of the Head of Hylebos dredging project. All the
water entrained with the dredged sediment was placed in the barge
rather than being released back to the water. Overall, the enclosed mechanical buckets placed more water than sediment in the barges. This
water can contain an important load of sediments and contaminants.
During the 2005 season, the water-management system captured about
4,000 cy of sediment. It is estimated that if that material had been released back to the dredge area, it would have generated a layer of impacted sediment an average of 3-4 in. thick over the dredged area. Capture of the solids contributed to the ability to meet the cleanup goals
(Dalton, Olmsted & Fuglevand, Inc. 2006).
Remediation contractors for the Todd Shipyard found benefit in
working with dredging companies that were able to mobilize an array of
equipment from their inventory to meet project needs and respond to
changing site conditions and schedules. The project engineer for Todd
Shipyard included the dredging contractor as a consultant during the
design stages of the remediation—an important step that ensured a
smooth transition from design to implementation (EPA 2006a [Todd
Shipyards Sediment OU, May 12, 2006]). Experienced environmental
dredgers also have the capabilities to operate within the typical regulatory restrictions and an understanding of the difficulties associated with
environmental dredging. The importance of using contractors experienced in environmental dredging was emphasized by the remediation
contractors at the Fox River SMU 56/57 dredging project:
Most large dredging contractors in the United States have little or
no experience with contaminated sediment projects, working pre-
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dominantly on navigational dredging projects. Navigational dredging projects typically have no environmental controls, resulting in
higher production rates and lower unit costs. Larger-scale projects
may also limit the available temporary water treatment and dewatering equipment unless planned well in advance, as well as onshore land space, that are necessary to complete the work in a
timely fashion (Montgomery Watson 2001).25
Since 1999, the first year of the SMU 56/57 dredging project, the
numbers and experience of firms experienced with environmental
dredging has increased. However, the need for contractors familiar with
the challenges of environmental dredging remains, particularly at large,
multi-year megasite projects.
Use Appropriate Contracting Arrangements
The nature of the contracting vehicle used to conduct the work can
drive behavior of the contractor and ultimately impact project results.
The contracting terms and approaches provide incentives for contractor
performance so the contracting approach needs to be aligned with the
project’s risk reduction goals. For example, the implementation of BMPs
can be encouraged or discouraged by the contracting mechanisms used
in a remedial project.
EPA described the importance of the contracting mechanism at the
Lockheed Shipyard cleanup (EPA 2006a [Lockheed Shipyard Sediment
OU, May 11, 2006]):
The primary keys to success for a complex project such as the
LSSOU cleanup include (1) a contract where the dredging contractor was not taking the risk and 2) the dredger was guaranteed a
daily rate for each activity. Frequently, in the past, environmental
dredging contractors have followed the “navigational dredging
It is true that navigation projects often have less environmental controls than
environmental dredging projects, but navigation dredging projects typically
have to comply with water quality certification and disposal requirements of the
Clean Water Act or, in the case of ocean disposal, the Marine Protection, Research, and Sanctuaries Act.
25
158
model” in terms of contract style and dredging methods. Under a
Unit Rate contract the dredger is in a "production dredging mode,
which does not work for environmental dredging. Under the Time
and Material [contract] the dredger is not penalized for taking the
appropriate time to accomplish the task at hand. Because of this the
dredger is more likely to comply with BMPs and to take care to
minimize loose of material back into the waterway or to cause resuspension.
Phase 1 dredging at the site resulted in incomplete debris removal,
the presence of undredged inventory, and a large amount of residuals at
the end of the in-water work window. On review of the approach, the
remediation project manager revised the contract mechanism for phase 2
dredging (second season) to use a time-and-materials approach and requested new bids for the work (EPA 2006a [Lockheed Shipyard Sediment OU, May 11, 2006]).
The Todd Shipyard project engineer found that a cost-plusincentive-fee form of contract worked well by motivating contractors to
complete every aspect of the construction in accordance with defined
quality objectives at the lowest overall cost. The contract reimbursed the
contractor for all direct costs of the work. That removed the financial risk
to the contractor and thereby reduced bid costs to cover unknown or unquantifiable risks (EPA 2006a [Harbor Island Todd Shipyards Sediment
OU, May 12, 2006]).
In the Head of Hylebos dredging project, a cost-plus-fee contract
assisted in the reduction of residual contamination. The engineers and
contractors determined that it was in their best interest to minimize the
extent of the residual layer by slowing down the production to match
“good housekeeping” on the bottom. That contracting mechanism provided the opportunity for the owner to work with the contractor to adjust construction and operations to achieve the project objectives. Contractor oversight was provided by onsite dredge inspectors during each
shift. That oversight was used to make decisions regarding whether the
target elevations were met (that is, finding the underlying native material) before dredging moved to the next cut. The cab on each dredge was
actually expanded to provide a place for the dredge observer to sit side
by side with the operator day and night throughout the dredging operation (Dalton, Olmsted & Fuglevand, Inc. 2006).
159
Overall, there is no single best contract type. These decisions will
depend on site conditions and necessary equipment and materials.
However, contracting terms and approaches that encourage contractors
to focus on achieving cleanup goals and remedial action objectives are
best suited for environmental dredging. These arrangements should create incentives for reducing resuspension and residual production, using
best management practices, and adjusting the dredging approach to improve chances of meeting cleanup levels and result in cost savings.
Use Operational Controls to Improve Dredging Accuracy
In addition to appropriate design, implementation, and monitoring,
technologic approaches can improve dredging efficiency and effectiveness. Some of them have been used successfully at contaminated sediment sites.
A one-of-a-kind specially designed high-technology dredge outfitted with innovative sensors and controls to achieve a 6-in. excavation-cut
tolerance was used to extract creosote-contaminated sediment at Bayou
Bonfouca, LA. A cutline to the depth associated with a total PAH concentration of 1,300 ppm was established and programmed as an absolute
elevation along a 4,000-ft length by using borehole concentration profiles.
Maximum contamination depth was 17 ft (average, 10 ft). Logs, concrete,
metal objects, and so on were removed with grapple hooks before excavation (EPA 2006a [Bayou Bonfouca Superfund site, May 12, 2006]). The
pre-dredging operation and low tidal fluctuations and low stream flow
rate (13 ft3/sec) provided a stable dredging platform (spud barge) in the
loose, high-organic-matter layer over “harder” unconsolidated inorganic
substrate, which all aided in the implementation of the new precision
dredging technology. No post-dredging measurements (such as bottom
elevations) were taken to evaluate achievement of the bottom cut-line
target programmed into the excavator, nor were any sediment analyses
performed to verify achievement of a total PAH concentration of less
than 1,300 ppm. Targets for volume of dredged material were achieved,
however, and that was the primary goal of controlling the excavation
depth.
The dredge used at Todd Shipyards was equipped with a positioning system with 20-cm (GPS-controlled) horizontal accuracy. It provided
160
real-time display and tracking of the horizontal and vertical position of
the dredge bucket. Digital GPS receivers and a gyrocompass were used
to determine real-time horizontal (X and Y) positioning of the derrick
barge and the dredge bucket. An electronic tide gauge was used to allow
the operator to determine the proper dredge elevation below the water
surface accurately. The vertical position of the bucket was combined
with the electronic tide-gauge data to determine the bucket elevation (Z).
In addition, the dredge-bucket wires were painted in 1-ft increments to
provide for a check on the electronically calculated vertical position. The
information generated by the positioning system was electronically
stored and used to create maps that showed dredging progress, including the degree of overlap between bucket deployments (EPA 2006a
[Harbor Island Todd Shipyards Sediment OU, May 12, 2006]). That approach to navigation and positioning of the dredge bucket has been used
at a number of sites in Puget Sound.
Precision positioning systems are seeing increasing use. Although
the basic technology of dredging has changed little in recent decades, the
ability to position the dredge accurately has improved dramatically; in
principle, this can improve our ability to remove contaminated sediment
accurately and efficiently if sufficient site-characterization data are available. Regardless of the improvements in dredging methods and equipment, the reliability of the equipment and the availability of skilled operators capable of processing and interpreting the data remain challenges.
Consider Backfilling and Capping to Control Residuals
As indicated previously, the factor limiting dredging effectiveness
that is the most difficult to manage is high residual contaminant concentrations. Residuals are always detected after dredging and can be relatively high in concentration and typically of the same order as the average concentration in the dredged material (Reible et al. 2003). The
magnitude of residuals can be higher in the presence of debris or when
site conditions make it infeasible to overdredge into clean material. Even
in favorable dredging conditions, however, some degree of residual control is usually necessary to achieve site cleanup standards and to address
site remedial action objectives. Generally, control of residuals is achieved
by adding backfill or thin-layer capping; this has clear advantages in
161
achieving bulk sediment contaminant concentration targets even if the
backfill layer is intermixed with the residual sediments. Although backfill can effectively manage bulk sediment concentrations, the effectiveness of backfill for aiding long-term risk reduction is less well understood. In addition, the advantages of dredging with backfill for residual
control relative to a complete capping remedy need to be assessed during remedy evaluation.
At Bayou Bonfouca in Slidell, LA, backfilling was a necessary phase
of the overall remedial operation. Excavation to up to 3 m compromised
the stability of the unconsolidated material forming the banks along the
bayou. Gravel (about 1 ft) over sand (about 1 ft) and additional fill gave
support to the sheet-piling-reinforced banks (5,000 ft long) and served to
cap the residual PAH contamination of 1,300 ppm (target concentration).
Backfilling to the original bottom grade maintained the historical water
flow rates and levels needed for recreational and other boating traffic
and allowed the bayou to begin natural recovery.
The remediation of the former Ketchikan Pulp Company facility, in
Ward Cove in Ketchikan, AK, also involved some backfilling of dredged
areas. The facility operated as a dissolving sulfite pulp mill from 1954
until 1997 and discharged untreated sulfite waste liquor (magnesium
bisulfite), pulping solids, and bleaching waste (chlorine caustic) into
Ward Cove until 1971, with increasing wastewater treatment after that.
Mill operations affected sediment by releasing large quantities of organic
material (up to 10 ft thick) as byproducts of wood pulping. The organic
material altered the physical structure and chemistry of the sediments
and thus the type and abundance of benthic organisms. Degradation of
the organic-rich pulping byproduct led to anaerobic conditions in the
sediment and production of ammonia, sulfide, and 4-methylphenol in
quantities that were potentially toxic to benthic organisms (EPA 2006a,
Ketchikan Pulp Company; April 26, 2006). Remedial action objectives
included reducing toxicity of surface sediments to benthic life and enhancing benthic recolonization. The selected remedy included thin-layer
(6-12 in.) placement of clean sand over dredged areas (to less than the
full depth of contamination) and undredged areas and monitoring of
natural recovery where thin-layer placement was not practicable. Sand
backfilling was expected to achieve the remedial objectives by diluting
contaminants and organic matter, both of which are associated with benthic toxicity on this site (EPA 2000b).
162
Remediation was completed in 2001. The long-term monitoring
program includes sediment chemical analysis, toxicity testing, and assessment of the benthic macroinvertebrate community. The first round of
long-term monitoring in 2004 found that the remedy appeared to have
met its objectives in backfilled areas; chemical concentrations were generally below sediment cleanup levels as determined from preremediation toxicity testing, survival of benthic test organisms was high,
and benthic species diversity and abundance increased relative to the
pre-remediation baseline and are similar to reference areas (Exponent
2005). Additional monitoring is planned for 2007, and equally favorable
or improved results may lead to a reduction in required monitoring of
backfilled areas. In contrast, 2004 monitoring results showed that only
one of four natural recovery areas had improvements comparable to
those found in the backfilled areas (Exponent 2005; Herrenkohl et al.
2006).
Backfilling has been used at a variety of sites, including the Fox
River and several sites in the Puget Sound area. It has been proposed for
many sites that have not met or are unlikely to meet cleanup levels after
dredging alone. Backfill that is at least about 6-12 in. thick probably
forms an effective separation between much of the benthic community
that might colonize the top of the backfill layer and the underlying sediment. Thin sand backfill layers (less than 6 in.), however, are of uncertain
effectiveness because the low sorptivity of sand means that the benthic
community may be exposed to pore water contaminant concentrations
similar to that of uncapped sediment. Exposure and risk are often more
closely related to pore water concentration than to bulk sediment concentration, and further research or field monitoring is needed to confirm
the appropriateness of thin-layer backfilling.
CONCLUSIONS
On the basis of its review of data and experiences at dredging projects, the committee has reached the following conclusions:
• The committee was generally unable to establish whether dredging alone is capable of achieving long-term risk reduction, because
o
163
Monitoring at most sites does not include the full array
of measures necessary to evaluate risk.
o
Dredging may have occurred in conjunction with other
remedies or natural processes, or insufficient time may
have passed to evaluate long-term risk reduction.
o
A systematic compilation of site data necessary to track
remedial effectiveness nationally is lacking.
• Dredging remains one of the few options available for the remediation of contaminated sediments and should be considered, with other
options, for managing the risks that they pose.
• Dredging is effective for removal of mass, but mass removal
alone may not achieve risk-based goals.
• Dredging will likely have at least short-term adverse effects on
the water column and biota.
• Dredging effectiveness is limited by resuspension and release of
contaminants during dredging and the generation or exposure of residual contamination by dredging. Those limitations are minimized if site
conditions are favorable and the remedy is designed and implemented
appropriately.
o
Favorable site conditions include
−
Little or no debris
−
A visual or physical texture difference or other
rapid mechanism for differentiating clean and contaminated sediments.
−
Potential for overdredging into clean material.
−
Low-gradient bottom and side slopes.
−
Lack of piers and other obstacles.
−
Site conditions that promote rapid natural attenuation after dredging (for example, through
natural deposition).
−
Absence of non-aqueous-phase liquid or readily
desorbable contaminants.
o
Effective design and implementation factors include
−
Site characterization sufficient to develop a
comprehensive conceptual site model and identify
adverse site conditions.
−
Identification and control of sources on a watershed-wide basis.
164
−
Use of pilot studies, where appropriate, to identify adverse site conditions and appropriate management responses.
−
Application of best management practices to
control residuals and resuspension (for example,
operational controls at the dredge and on produced
streams, appropriate equipment selection, and residual control measures).
−
Contracting and procurement mechanisms to
encourage a focus on cleanup levels and remedial
action objectives.
−
Engagement of experienced and innovative environmental-dredging contractors throughout the
design and implementation phases of remediation.
• Dredging alone is unlikely to be effective in reaching short-term
or long-term goals where sites exhibit one or more unfavorable conditions. Where unfavorable conditions exist, increased contaminant resuspension, release, and residual will tend to limit ability to meet cleanup
levels and delay the achievement of remedial action objectives unless
managed through a combination of remedies or alternative remedies.
RECOMMENDATIONS
• A remedy should be designed to meet long-term risk-reduction
goals. The design should be tested by modeling and monitoring the
achievement of long-term remedial action objectives.
• Site conditions that influence dredging effectiveness should be
recognized during selection, development, and implementation of the
remedy. When conditions unfavorable for dredging exist:
o
Implementation of one or more pilot tests should be considered to identify optimal remedial approaches and assess
their effectiveness.
o
Adverse effects of resuspension, release, and residuals
should be forecast and explicitly considered in expectations
of risk.
o
The ability of combination remedies to lessen the adverse effects of residuals should be considered when evaluating the
potential effectiveness of dredging.
165
o
Best management practices should be implemented to
minimize effects of adverse dredging conditions.
o
The possibility of adverse dredging conditions that are
not anticipated should be recognized and planned for.
• A good baseline assessment coupled with a well-designed longterm monitoring plan should be implemented to permit evaluation of
dredging effectiveness.
o
Well-designed pre-dredging and post-dredging monitoring is necessary to establish effectiveness and indicate
achievement of remedial action objectives.
o
Monitoring should be conducted to demonstrate
achievement of cleanup levels and to confirm that the
cleanup levels achieve remedial action objectives.
o
Data from monitoring should be managed and stored in
electronic databases accessible for further analysis.
• Further research, including during dredging pilots and full-scale
operations, should be conducted to define mechanisms, rates, causes,
and effects of dredging residuals and contaminant resuspension.
REFERENCES
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Available: http://www.dnr.state.wi.us/ORG/water/wm/foxriver/cuments/
sediment/usgs_5657_report.pdf [accessed Jan. 11, 2007].
Stow, C.A., S.R. Carpenter, L.A. Eby, J.F. Amrhein, and R.J. Hesselberg. 1995.
Evidence that PCBs are approaching stable concentrations in Lake Michigan fishes. Ecol. Appl. 5(1):248-260.
TetraTech. 2004. Final Interim Remedial Action Report for Site 1—McAllister
Point Landfill, Operable Unit 4—Marine Sediment/Management of Migration, Naval Station Newport, Middletown, Rhode Island. Prepared for Department of the Navy, Engineering Field Activity, Northeast Naval Facilities Engineering Command, Lester, PA, by Tetra Tech. FW, ICN,
Langhorne, PA. September 28, 2004.
TetraTech NUS. 2006. Round 1: December 2004 Long-Term Monitoring Report
for McAllister Point Landfill Naval Station Newport Newport, Rhode Island. Prepared for Engineering Field Activity Northeast Naval Facilities
Engineering Command, Lester, PA. Contract Number N62472-03-D-0057.
March 2006.
Thibodeaux, L.J., and K.T. Duckworth. 1999. The effectiveness of environmental
dredging: A study of three sites. Remediat. J. 11(3):5-33.
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URS (URS Woodward Clyde Group Consultants Inc.). 1999. Construction Monitoring Report Christina River Remediation: Newport Superfund Site,
Newport, DE. Prepared for DuPont Corporate Remediation Group, Wilmington, Delaware, by URS Woodward Clyde Group Consultants Inc.,
Buffalo, New York. December 1999.
URS (URS Group Inc.). 2006. Final 2003 Marine Monitoring Report, OU B Marine,
Bremerton Naval Complex, Bremerton, Washington. Prepared for Naval
Facilities Engineering Command Northwest, Poulsbo, WA, by URS Group,
Inc., Seattle, WA. February, 2, 2006.
USACE (U.S. Army Corps of Engineers). 1992. Remediation of Marathon Battery
Superfund Project: Cold Spring, New York, Supplemental Information. IFB
No. DACW41-92-B-0030. U.S. Army Corps of Engineers, Kansas City District. May 1992.
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Dredge Sediment Sampling. Memorandum to Dave Dickerson, U.S. Environmental Protection Agency, from Mark Otis, U.S. Army Corps of Engineers. CENED-PD-E. April 19, 1995.
USACE (U.S. Army Corps of Engineers). 2005. Benthic Recovery Review. Draft
Technical Note ERDC TN-DOER-E. August 2005.
Weston (Weston Solutions Inc.). 2002. Final Comprehensive Post-Removal Summary Report Manistique Harbor and River Site U.S. Route 2 Highway
Bridge Manistique, Michigan. Prepared for the U.S. Environmental Protection Agency, Region 5. November 12, 2002.
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Manistique, Michigan, Data Evaluation Report: Revision 1-19 May 2005.
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190/219190_HHRA_Vol1.pdf [accessed May 7, 2007].
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and Operable Unit 2, Lower Fox River and Green Bay. Wisconsin Department of Natural Resources, Madison, WI, and the U.S. Environmental Protection Agency, Region 5, Chicago, IL. December 2002 [online]. Available:
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ecord_of_Decision_OUs_1-2.pdf [accessed Jan. 11, 2007].
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Windward. 2005. Lower Duwamish Waterway Remedial Investigation. Data
Report: Fish and Crab Tissue Collection and Chemical Analyses. Final Report. Prepared for U.S. Environmental Protection Agency, Region 10, Seattle WA, and Washington State Department of Ecology, Bellevue, WA, by
Windward Environmental LLC, Seattle, WA. July 27, 2005 [online]. Available:
http://www.ldwg.org/assets/fish_crab_tissue/final_fish%2bcrab_
tissue_dr.pdf [accessed April 25, 2007].
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Report: Chemical Analyses of Fish and Crab Tissue Samples Collected in
2005. Final Report. Prepared for U.S. Environmental Protection Agency,
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Bellevue, WA, by Windward Environmental LLC, Seattle, WA. April 19,
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dr/2005_fish%2bcrab_dr.pdf [accessed April 25, 2007].
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php/publications/html/ecolsed/ [accessed Sept. 12, 2006].
5
Monitoring for Effectiveness: Current
Practices and Proposed Improvements
MONITORING FOR EFFECTIVENESS
The effectiveness of environmental dredging in reducing risk, as
predicted when the remedy was selected, can be verified only through
monitoring. Monitoring includes
• Monitoring of potential short-term risks due to dredging.
• Verification that dredging has achieved its immediate target
cleanup levels.
• Long-term monitoring to determine whether remedial objectives
have been or are likely to be achieved in the expected time frame.
Monitoring of effectiveness is an essential part of the remedy and
should be proportional to the size of the project. Through careful monitoring it is possible to demonstrate whether environmental dredging
minimizes risks to human and ecologic receptors during active operations and to judge the success of contaminant cleanup in decreasing risk
after the cessation of active remedial operations. Monitoring is the only
way to determine short-term and long-term compliance with remedialaction objectives and evaluate net risk reduction of the remediation, and
it forms the basis of the 5-year performance reviews after cleanup. Be178
Monitoring for Effectiveness
179
cause sediments typically pose long-term risks, monitoring often must
span decades to assess risk reduction.
The ultimate goal of monitoring is protection—that is, ensuring that
short-term and long-term risks are minimized, by providing sufficient
information to judge that the remedy is effective, or to adapt site management to optimize the remedy’s performance to achieve risk-based
objectives. Management adaptation may entail modification of dredging
procedures—for example, if short-term exposures exceed expected magnitudes—or modification of the remedy itself by amendment or modification of the record of decision (ROD) if long-term risk reduction proceeds more slowly or more rapidly than expected.
An effective sediment-monitoring plan takes into account the successive stages of sediment cleanup: site characterization; selection, planning, and implementation of the remedial action; effectiveness assessment; and adaptive management.1 Monitoring should build on the
studies previously performed for the remedial investigation and feasibility study (RI/FS), which should have
• Determined the nature and extent of contamination and any
trends in time (for example, due to natural recovery).
• Supported or developed a conceptual site model.
• Provided information to assess risks to the environment and
people.
• Evaluated remedial alternatives, including a quantitative comparison of risks associated with implementation of each one.
Once the remedy is selected and implementation begins, monitoring extends the record of site conditions into the future.
MONITORING PRINCIPLES
1. Monitoring should be based on and inform the conceptual site
model.
In general, adaptive management is the testing of hypotheses and conclusions and re-evaluation of site assumptions and decisions as new information is
gathered (see Chapter 6 for further discussion). It is an important component of
the updating of the conceptual site model (EPA 2005a).
1
180
a.
Appropriate metrics need to be chosen, measuring success against expectations based on the conceptual site
model.
b. Monitoring is an essential verification step, not an addon activity or a second remedial investigation.
2. Effective monitoring of the remedy requires characterization of
pre-remedial trends and reference conditions, in addition to postremedial trends.
a. Sufficient pre- and post-remedial sample sizes are
needed, to allow for natural heterogeneity.
b. The time span of pre- and post-remedial sampling
needs to be sufficient to capture the time scale of recovery processes.
c. Proper reference sites and conditions must be specified
and monitored.
Monitoring and the Conceptual Site Model
Links between contamination, exposure, and risk can be highly
complex, involving multiple physical, chemical, and biologic processes.
A particular combination of these is present at each site. Monitoring protocols and media to be monitored will vary accordingly, and should be
closely linked to site conceptual models that link site conditions with
biologic exposures and effects (EPA 1998). The expectations of the Superfund ROD are a natural yardstick against which to judge effectiveness.
Those expectations of short-term exposures and long-term risk reduction
due to dredging should be based on the conceptual site model and its
mathematical counterpart.
Where site conceptual models are insufficiently developed, it is difficult to develop an understanding of the factors driving trends in sitemonitoring data. On major dredging sites, short-term and long-term expectations based on site models will have been developed as part of the
feasibility study supporting remedy selection. Collecting data to test
whether expectations have been fulfilled is part of the process of conceptual-model development, testing, and refinement that was begun with
the initial site characterization. If the important cause-effect relationships
between contaminant sources, transport mechanisms, exposure path-
181
ways, and receptors have been well characterized by the time a remedy
is selected, including bioavailability and food-web relationships as applicable, and there has been sufficient pilot testing or other means of anticipating site-specific field conditions and implementation challenges,
well-designed monitoring should indicate the remedy has performed as
expected. If not, monitoring can help to identify important elements that
are missing from the conceptual site model so that its predictions can be
made more accurate and site management can be adapted accordingly,
as recommended in EPA’s Contaminated Sediment Remediation Guidance
(2005a).
In monitoring of the effectiveness of a remedy, important transport
mechanisms and exposure pathways to be monitored include not only
the ones that control exposures and risks under normal conditions, but
also the ones that may be triggered by dredging, such as releases that
may occur during normal dredging operations or when debris or bedrock is encountered. Therefore, before selection and implementation of a
remedy, the site investigation should thoroughly examine factors that
would complicate dredging and include them in the conceptual model.
Complicating site conditions and operational limitations can also be
identified through pilot studies to verify the performance of the selected
technology under site-specific conditions.
Data collection is one of the more expensive aspects of site management (Box 5-1).2 Judicious use of the conceptual site model in designing the monitoring plan focuses data collection where it can best ensure
protectiveness while conserving monitoring resources. Monitoring
should target the key pathways and receptors necessary to determine
whether remedial objectives have been met. If dredging is intended to
reduce ecologic or human health risks, the conceptual site model can be
used to focus sampling on locations and receptors that directly indicate
risk related to the targeted sediments and contaminants and minimize
spurious effects, such as increased body burdens in migratory species
In addition to the example provided in Box 5-1, see the breakdown of costs of
the Hylebos Waterway and 2004 dredging at New Bedford Harbor presented in
Chapter 2 (Figures 2-4 and 2-5). However, it should be noted that these costs may
not be directly comparable; it is not clear, for example, whether the costs include
design costs and long-term monitoring.
2
182
BOX 5-1 Estimated Monitoring Costs for Lower Fox River and
Green Bay, Wisconsin, ROD Remedy
The costs of construction monitoring (including verification sampling) and
long-term monitoring (including an initial pre-dredging baseline survey of affected media and surveys of the same media continuing for decades after the
remedy) for the ROD remedy for Operable Units 2-5 of the Lower Fox River and
Green Bay, WI, have been estimated at $6 and $8 million, respectively (Shaw
2006). Together those costs exceed the estimated cost of engineering and construction support for the remedy, including development of design documentation, plans, and specifications.
and species with wide home ranges that are due to unrelated exposures
at remote locations. If dredging is intended to minimize water-column
contaminant transport, the site model can be used to control for the effects of flow, temperature, seasonality, non-sediment-related stressors
(such as point and nonpoint sources), and other ambient conditions to
inform sampling plans and assist in interpreting the results.
Monitoring decisions may be influenced by financial, jurisdictional,
or political interests, even though they should be guided solely by the
need to verify conceptual site models, inform remedy implementation,
and to document when remedial objectives have been achieved. Cleanup
negotiations between regulators and responsible parties can be contentious, and agreements on the scope of cleanups are often the results of a
long and difficult process. The scope of post-remedial monitoring can
also be established during those negotiations. The parties have few incentives to seek actively to establish whether a chosen remedial action
had its intended effect. This paradigm, wherein both regulators and responsible parties may perceive that they have something to lose and
nothing to gain in a robust post-remediation monitoring program, may
be a reason for the lack of post-remediation confirmation sampling seen
at some sites. Public-sector and private-sector designers of a monitoring
plan may face strong pressures to demonstrate early success while controlling costs and may also feel pressure to divert remedial funding to
support broader long-term natural-resource monitoring efforts. Those
ancillary goals may be attractive to parties involved in designing a monitoring program, but the fundamental objectives of monitoring are to per-
183
form a fair and conclusive evaluation of remedy effectiveness and risk
reduction, and resources and energy should be focused on this objective.
Information developed from the monitoring program should be used to
guide future decision-making in a manner which balances a realistic assessment of the projected environmental benefit relative to anticipated
costs.
Developing a body of well-designed site evaluations of dredging
effectiveness will meet the broader programmatic objective of providing
EPA and other lead agencies with invaluable information on strengths
and weaknesses of dredging as a remedy—information that they can use
in future remedial decision-making.
Comparisons to Baseline Conditions
To assess the effectiveness of the remedy, post-remedial monitoring
should be compared with data trends and model forecasts developed
before remedy selection. This requires that there be comparable datasets
before (a “baseline”) and after dredging. As stated by EPA (2005a, page
8-2),
During site characterization, the project manager should anticipate
expected post-remedy monitoring needs to ensure that adequate
baseline data are collected to allow comparison of future datasets.
Monitoring plans should also be designed to allow comparison of
results with model predictions that supported remedy selection.
It is often difficult in practice for an effective monitoring plan to
meet the above objectives. One important issue at Superfund megasites
is that the time from initial site investigation to implementation of remedial measures can be 10 years or more; it is extremely difficult to ensure
temporal and spatial consistency of baseline and post-remedial monitoring data, including data- quality assurance and control. Data collections
that span many years can greatly complicate the selection of appropriate
statistical tests for evaluating them. Those concerns are often manifested
after the fact rather than being evident during the planning of the baseline and long-term monitoring programs.
184
Consistent with its role in supporting hypothesis-testing, the monitoring protocol should be rigorous enough to allow managers to evaluate
critically the potential adverse effects of dredging on human and ecologic receptors and potential risk reductions due to removal of contaminated sediment. For example, proper reference sites or reference conditions should be established to allow comparison of affected media with
pre-dredging or nondredged controls. Appropriate sample sizes should
be determined from estimates of variability derived from pilot studies or
other sources of data. In particular, the natural heterogeneity of biologic
systems can be substantial and should be explicitly accounted for in defining sample sizes.
CURRENT MONITORING PRACTICES
According to Elzinga et al. (1998, as referenced in EPA 2004), monitoring is “the collection and analysis of repeated observations or measurements to evaluate changes in condition and progress toward meeting
a management objective.” Monitoring at Superfund sites is typically directed toward evaluation of the performance of a remedy and whatever
environmental protections are in place during implementation of the
remedy. Monitoring may include the collection of samples or real-time
metered data
• During implementation of the remedy to assess immediate human health or environmental effects.
• Soon after implementation to determine compliance with
cleanup levels or other short-term objectives.
• Over time to evaluate the achievement of the long-term remedial-action objectives, the need for maintenance or repair, and the continued effectiveness of the remedy and associated source control.
Ideally, the monitoring parameters measured are linked to sitespecific risk factors so that success (or lack of success) of the remedy is
evident and directly informs management of the site. There are no absolute requirements for monitoring elements or techniques, but a number
of guidance documents have been published (Fredette et al. 1990;
EPA/USACE 1998; EPA 2001a, 2004, 2005a) to identify relevant meas-
185
urements and techniques and to guide the design of monitoring programs for a contaminated sediment site undergoing remediation.
Monitoring Parameters and Techniques
Monitoring involves combinations of physical, chemical, and biologic methods. Three critical lines of evidence that increasingly define
successful sediment remediation include sediment physical stability,
sediment chemical stability (lack of movement of contaminants from the
sediment to the water column), and biologic-ecologic integrity. These
three concepts are integral components of remedy evaluation, and monitoring should use techniques sufficient to measure progress toward these
end points.
A variety of techniques and measurement parameters exist for the
characterization of the nature, extent, and potential effects of sediments.
These techniques range from relatively simple and quick to elaborate
and time consuming (e.g., EPA 2001a; Wenning et al. 2005). Several of
the techniques are described below and summarized in Box 5-2.
Physical Techniques
Available physical techniques include direct sampling of sediment
for laboratory analysis of geophysical properties, core sampling to identify sediment layering or the presence of debris, side scan sonar to develop high resolution maps of bottom contours, acoustic sub-bottom profilers or magnetometers to map sub-bottom characteristics, remote
sensing to document vegetative cover or other characteristics, videography or photography to document bottom features or shallow sediment
profile characteristics, and instrumentation to measure environmental
conditions (such as temperature and turbidity) or flow characteristics
that may affect sediment and suspended solids transport. For example,
sediment-profile imaging (a photographic technique) of surface (10-20
cm) characteristics can be conducted to establish various parameters including the depth of bioturbation, the depth of an oxygenated layer,
general benthic community type and degree of recovery, or hydrogen
sulfide gas production (see Figure 5-1 for an example). Other remote
186
BOX 5-2 Common Physical, Chemical, and Biologic Measurements Used
To Characterize Contaminated Sediments
Common physical measurements include
• Sediment geophysical properties, such as bulk density, particle size, and
shear strength.
• Pre-dredging and post-dredging bottom elevations, and sediment bedforms.
• Sediment layering, such as depth of disturbance or bioturbation, presence
of gas bubbles, redox layers, and interfaces between sediment of different textures.
• Debris-field mapping (location, density, and size).
• Conductivity, temperature, turbidity, and suspended particles under
various flow conditions.
• Stream velocities.
Common chemical measurements include
• Water-column parameters (such as dissolved oxygen and total and dissolved chemicals under various flow conditions).
• Surface- and subsurface -sediment chemistry, including magnitude, distribution, and depth of contamination.
• Pore water contaminant concentrations.
• Bioavailable fractions of contaminants in sediment, on the basis of organic-carbon normalization or acid volatile sulfide (AVS) analysis.
• Tissue contaminant concentrations including tissues ingested by humans
(in field collected or exposed aquatic organisms or plants) or tissue surrogates.
• Air quality (including odor) during construction of remedy or handling
of dredged material.
Common biologic measurements include
• Benthic invertebrate community structure (including abundance, diversity, and other structural or functional indexes).
• Toxicity (acute and chronic effects measured in the laboratory or field).
• Aquatic or wetland plant community structure (including species composition and percentage of cover).
• Fisheries status (including size, abundance, reproductive status, and incidence of lesions or parasites).
187
FIGURE 5-1 Sediment profile imagery (SPI) equipment (two left photos) and
sediment profile photograph (right) from New Bedford Harbor Superfund site
(the outer harbor is the area of the site with the least contamination). This
equipment is used as part of the long-term monitoring program at the site to
assess benthic quality rapidly and augment traditional benthic survey techniques
that entail sieving and enumeration. Source: W. Nelson, U.S. Environmental Protection Agency.
sensing techniques, such as lidar (light detection and ranging) can be
used to map large-scale site characteristics, including the extent of eel
grass beds or other vegetative cover. Assessments of physical stability of
sediments (which translates into the likelihood for sediments to be dislodged and transported by erosive events) are based on site uses, hydrology and geomorphology, sediment bed descriptions (radio dating
deposits, stratigraphy, and physical characteristics), and measurement of
sediment transport and sediment bed dynamics (erodability or bed elevation changes) (Bohlen and Erickson 2006).
Chemical Monitoring
Chemical monitoring can address multiple media⎯including air,
sediment, water, biota, groundwater, and pore water⎯and can be designed to evaluate specific phases of chemicals of concern (for example,
if they are dissolved or suspended in association with solids). It is impor-
188
tant to monitor those parameters that affect chemical bioavailability,
such as total and dissolved organic carbon, acid volatile sulfides (AVSs),
grain size, and pore water fractions because organisms are exposed only
to the bioavailable fraction (NRC 2003). The relationship between chemical concentrations, the bioavailable fraction, and toxic effects is the foundation for establishing sediment quality guidelines (see next section).
Chemical sampling may involve in situ instrumentation for water, single-point grab samples of water or sediment obtained with various devices, or use of samplers that integrate chemistry over time or space
(such as sediment traps, composite water samplers, and peepers). Rapid
chemical screening techniques that use immunoassay response (enzymelinked immunosorbent assays [ELISA]) or chemical fluorescence to
document relative exposures have also been developed, but these are
generally single-contaminant or contaminant-class tests, and few rapid
field screening techniques are available for measuring a broad array of
contaminants.
Some analytic methods for environmental samples can be timeconsuming, labor-intensive, and expensive. For example, chemical
measurements for persistent organic contaminants in sediments—such
as polychlorinated biphenyls (PCBs), polycyclic aromatic hydrocarbons
(PAHs), and DDT—require extraction, cleanup, and instrument analyses
with gas chromatography or mass spectrometry. None of those measurements is rapid or performed conveniently in the field. As with sediments, the chemical analysis of biologic samples requires extraction,
cleanup, and instrument techniques. Replicate measurements are necessary for both sediment samples and biologic tests because of inherent
variability.
Newer techniques have been developed for deployment of manufactured materials in the form of passive sampling devices (such as
semipermeable-membrane devices, solid-phase microextraction fibers,
and Tenax) that can mimic biologic exposure to and tissue uptake of contaminants from water, pore water, or sediments. For example, semipermeable membrane devices (SPMDs) have been widely used in environmental applications since the early 1990s (Huckins et al. 1990; 1993) and
applied at Superfund sites to monitor dissolved hydrophobic contaminants and estimate water column concentrations of these contaminants
(e.g., Hofelt and Shea 1997; Weston 2005). Polyethylene devices (PEDs)
passively sample hydrophobic organic compounds in the aqueous phase.
189
They are robust, simple, and inexpensive and have a short equilibration
time (Booij et al. 1998; Adams 2003). The laboratory analysis of PEDs has
fewer background-signal problems than the analysis of biologic samples.
Another example is solid-phase microextraction (SPME) of sediment
pore water, which may be quicker and more economical than conventional sampling. SPME uses fibers coated with a liquid polymer, a solid
sorbent, or a combination. The fiber coating removes the compound from
solution by sorption. The SPME fiber is then inserted directly into a gas
chromatograph for desorption and analysis.
Some of these passive sampling devices can reach equilibrium with
environmental conditions much more rapidly than living organisms.
They have been widely used in recent years for sampling metals and organics in aquatic systems and found to be good indicators of fish and
invertebrate bioaccumulation, that is, to be biomimetic (e.g., Arthur and
Pawliszyn 1990; Huckins et al. 1993; Wells and Lanno 2001). Biomimetic
samplers are easy to deploy and analyze, and they indicate exposure
over time, but they are selective and do not indicate all chemical exposures or biologic effects (Table 5-1). Several of these techniques are still
undergoing development and research is being conducted to better understand the relationship between the sampler concentrations, environmental concentrations, and bioaccumulation in organisms (e.g., Leslie et
al. 2002; Lohmann et al. 2004; Vinturella et al. 2004; Conder and LaPoint
2005; You et al. 2006). With time and refinement, these technologies will
likely become more available for routine application at contaminated
sediment sites.
Biologic Monitoring
Biologic monitoring looks at the sublethal to lethal responses of individual organisms, populations, or communities in the environment or
under controlled laboratory conditions. Biologic measurements and end
points are usually more complex or difficult to obtain than physical or
chemical measures, but biologic monitoring is the most definitive way to
determine risk. Biologic monitoring typically provides a more integrated
measurement of exposure (of both human and ecologic receptors) and is
related more directly to ecosystem effects than is physical or chemical
monitoring.
190
TABLE 5-1 Strengths and Limitations of Methods for Assessing Biologic Effects
in Aquatic Ecosystems
Effect Assessment
Method
Advantages
Limitations
Criteria or
Proven utility and ease of use
Assume single chemical
effect; based on laboratory
exposures; causality link
uncertain
guidelines
Biotic-ligand
model
Insufficient research and
for accounting for metal
validation for use with
bioavailability in surface water sediment
Empirically
based guidelines
Bioavailability not accounted
for; may lead to incorrect
conclusion of presence or
absence of risk
Equilibriumbased guidelines
Regulatory support; predictive
capability
Not applicable in dynamic
systems; does not consider all
critical binding phases
Species sensitivity
distributions
Use of all available data for
derivation of EQC or PNEC
Lack of sufficiently large and
diverse sediment-toxicity
datasets
Indigenous biota
Target receptors; lack of
laboratory extrapolation; longterm measure; proven utility;
public interest; colonization
and transplant methods
increase stressor diagnostic
power and experimental
power
Habitat and other natural
stressors or linkages
confound causality linkage;
inherent variability; loss of
colonization units possible
because of flow and
vandalism
Tissue residues and Documents exposure; use for
biomarkers
food web and risk models;
widely used; very sensitive
and timely
Adaptation, acclimation, and
metabolism confound
interpretations; uncertain
adverse- effect threshold
levels
Biomimetic
devices:
semipermeable
membrane devices;
Selectivity varies with
different chemicals; may not
mimic bioaccumulation of all
organisms; some are subject
Accumulates organics or
metals from waters and
sediments through diffusion
and sorption; amounts
191
TABLE 5-1 Continued
Effect Assessment
Method
Advantages
Limitations
solid-phase
microextraction;
Tenax; diffusive
gradient transport
accumulated on these inert
materials are similar to
amounts bioaccumulated in
fish tissues; can be placed in
situ for short to long periods
and then directly analyzed in
laboratory
to fouling, depending on
ecosystem; not standardized
Toxicity assays
(laboratory)
Bioavailability indicator;
proven utility; integrates
effects of multiple chemicals;
does not measure natural
stressors
Causality link uncertain;
laboratory-to-field
extrapolation; individual-tocommunity extrapolations;
does not measure natural
stressors; cost of chronic
assays
Toxicity and
bioaccumulation
assays (field)
More realistic exposure, which
reduces artifact potential;
measures many natural
stressors and interactions;
compartmentalizes exposures
to various media; exposure-toeffect linkage is strong
Most methods are not
standardized; limited use;
deployment can be difficult;
possible caging effects with
some organisms; causality
link uncertain; loss of units
possible because of predators
and vandalism; acclimation
stress possible because of
temperature, salinity, or
hardness differences
Toxicity
fractionation
(laboratory)
Better establishes specific
chemical causality; standard
method for effluents
Subject to manipulation
artifacts; acute toxicity only;
limited use in sediments;
large pore water volume
requirements; limited
sensitivity
Toxicity
More realistic exposure, which
fractionation (field) reduces artifact potential;
better establishes chemical
causality
Source: Modified from Burton et al. 2005.
Very limited use; deployment
can be difficult; shallow
environments only; acute
toxicity only; loss of units
possible because of high flow
and vandalism; not
standardized
192
Biologic testing often has both field and laboratory components—
organisms collected from the field are identified and enumerated or, in
marine systems, exposed to sediment-bound or water-borne chemicals in
a laboratory (e.g., Barbour et al. 1999; EPA 2001a; Adams et al. 2005). Indigenous organisms can be collected with nets, hooks, traps, grab samplers, or other devices. Standardized laboratory sediment toxicity and
bioaccumulation testing methods are commonly used in assessments of
the potential hazard of dredged materials (e.g., Environment Canada
1992; EPA 1994, 2000a; EPA/USACE 1998; ASTM 2006). Standardized
methods are available for freshwater and marine systems, in both short
and long term (chronic) exposures. These tests often are one component
of a “Sediment Quality Triad” and other weight-of-evidence based approaches (Adams et al. 2005). A strong relationship has been documented between the responses in these standardized laboratory test responses (and indigenous benthic communities) and empirically-based
sediment quality guidelines (discussed below) (Ingersoll et al. 2005).
Field-collected or laboratory-reared organisms can be deployed in
cages or nets for defined periods of exposure to water or sediment and
retrieved for analysis (e.g., Ireland et al. 1996; Tucker and Burton 1999;
Burton et al. 2000; Chappie and Burton 2000; Greenberg et al. 2002; Adams et al. 2005; Crane et al. 2007). These caged-organism assays allow
measurements of effects on growth and survival to be closely linked to
environmentally-relevant chemical exposures (Table 5-1) (Solomon et al.
1997; Burton and Pitt 2002; Adams et al. 2005; Wharfe et al. 2007). Transplantation and recolonization of benthic macroinvertebrates on reference
and site sediments have also been shown to be effective ways to measure
site effects and risk, but they require exposures of up to a month (e.g.,
Clements and Newman 2002; Clark and Clements 2006). Biologic monitoring can be cost-effective, relative to chemical monitoring (Karr 1993;
Hart 1994).
Transplantation, colonization, and caged-biota tests can be long
and have deployment challenges. However, caged exposures often take
only one to several days in freshwater systems (Ireland et al. 1996;
Tucker and Burton 1999; Chappie and Burton 2000; Burton et al. 2002,
2005; Greenberg et al. 2002; Burton and Nordstrom 2004). In situ caged
exposures of 2-4 days have been shown to provide uptake and toxicity
information that is comparable with that of standardized laboratory tests
that take 10-65 days at PCB-, chlorobenzene-, and metal-contaminated
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sediment sites (Greenberg et al. 2002; Burton et al. 2005). It is also useful
to conduct laboratory-based exposures following standardized toxicitytest methods (EPA 2001a).
Monitoring Human Exposures
The biologic monitoring techniques listed and discussed above are
related primarily to ecologic receptors and do not include monitoring of
human subjects, which EPA and other Superfund lead agencies do not
typically perform at contaminated sediment sites. However, biomonitoring of people who live near contaminated sediment sites is sometimes
performed by other parties, such as local health authorities and academic
scientists (Miller et al. 1991; Fitzgerald et al. 1996, 1999, 2004; MA DPH
1997; Korrick and Altshul 1998) and the Agency for Toxic Substances
Disease Registry.
Typical indicators of human exposure and risk reduction include
contaminant concentrations in the subset of environmental media to
which people might be directly or indirectly exposed in places where
exposures might occur. Those media include surface sediment and surface water in areas accessible to people and aquatic biota used for food.
Where there is interaction between contaminated sediments and floodplain, the list may be expanded to include floodplain surface soil, terrestrial game species foraging in the floodplain, and agricultural products
from the floodplain (see further discussion in sections below).
Use of Sediment Quality Guidelines
SQGs are numerical chemical concentrations intended to be either
protective of biologic resources or predictive of adverse effects to those
resources (Wenning and Ingersoll 2002). They are used to estimate the
toxicity and risk from sediments. At a contaminated site, the SQGs can
be used to establish contaminants of concern (COCs) from potentially
long lists of contaminants of potential concern (COPCs), identify or rank
problem reaches in a waterway, and classify hot spots (Long and MacDonald 1998).
There are two basic categories of SQGs, empirical and deterministic. Empirical approaches use statistical methods to compare sediment
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chemistry to effects datasets to predict the probability of or the presence
and absence of adverse or toxic effects (Word et al. 2005). A variety of
those approaches have been used to develop toxic effects levels, thresholds, or concentrations (used as SQGs) (MacDonald et al. 2000; Burton
2002; Wenning and Ingersoll 2002). Deterministic approaches typically
use equilibrium partitioning theory (Adams et al. 1985; Di Toro et al.
1991, 1992) to relate toxic concentrations found in water-only exposures
to sediment exposures for the same organism. Effects are predicted to
occur when toxic concentrations found in water occur in the pore water
of the sediment (Word et al. 2005); complexing agents (organic carbon
for hydrophobic non-ionic contaminants [Di Toro et al. 1991] and AVS
for cationic metals [Di Toro et al. 1990, 1992]) are the basis of the equilibrium calculations.
There has been a lot of controversy and discussion on the use and
viability of SQGs including their false positive and negative rates, their
applicability to mixtures of chemicals, their ability to establish cause and
effect relationships, and whether results can be extrapolated across species or biologic communities (Burton 2002; Wenning and Ingersoll 2002).
Sediment quality guidelines are only indirect measures of effects and do
not clearly establish whether risk or adverse biologic impacts are actually
occurring (Table 5-1) (NRC 2003).
A recent Pellston workshop summary, Use of Sediment Quality
Guidelines and Related Tools for Assessments of Contaminated Sediments
(Wenning et al. 2005), comprehensively reviews these approaches. A few
conclusions reached from this workshop include that (1) although the
scientific underpinnings of the different SQG approaches vary widely,
none of the approaches appear to be intrinsically flawed; (2) chemicallybased numeric SQGs can be effective for identifying concentration
ranges where adverse biologic effects are unlikely, uncertain, and highly
likely to occur, and; (3) in all cases, application of SQGs in a “toxic or
nontoxic” context must be cognizant of the types and rates of errors associated with each type of SQG (Wenning and Ingersoll 2002).
EPA has supported the development of mechanistically based
sediment quality guidelines (EPA 2000b, 2003a,b,c, 2005b) and the National Oceanographic and Atmospheric Administration (NOAA) has
supported the development of empirical sediment guidelines (Long and
Morgan 1990). It is expected that as the scientific issues continue to be
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resolved, they will see continued and greater use in toxicity evaluations,
comparative risk analyses, and in remedial decision making.
MONITORING-PROGRAM DESIGN
Selection of the appropriate monitoring measures and design of a
monitoring program depend on the development of clear hypotheses to
be tested or questions to be answered that are directly linked to a detailed conceptual site model characterizing sources, pathways of exposure, and receptors that may be exposed during or after remediation. By
the time a remedy is implemented, the understanding of site processes
should be highly refined so that monitoring can be focused on the expected beneficial and adverse effects of remediation. These effects include releases to the water column or atmosphere, as monitored during
dredging; residual sediment concentrations, as monitored by progress
samples or post-dredging verification sampling; and reductions in exposures and risks, as observed through long-term monitoring.
Monitoring During Dredging
Dredging operations include material removal, transport, dewatering, final disposal, and onsite solids treatment, water treatment, and
temporary storage. Therefore, monitoring during dredging operations
may involve a variety of activities, including some that are not directly
related to dredging operations or performance. An inherent difficulty is
the need for rapid measurement techniques that can inform contingency
actions and provide near-real-time feedback for executing corrective
measures while the work is ongoing.
Monitoring programs implemented during dredging are often
based on the requirements of Clean Water Act Rule 401 water quality
certification, typically administered by state environmental agencies. The
focus of any required monitoring for water quality certification is effects
on water quality, based on comparison with state or federal water quality standards and criteria, taking upstream conditions into account. Surrogate or indicator parameters (such as turbidity or concentrations of a
single chemical) are typically used to provide rapid information to the
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dredger and site manager and to develop a compliance history spanning
the various phases of the project.
With available technologies, some contaminant release and transport is inevitable during dredging (EPA 2005a). Depending on the volatility of the contaminant, there may be release to the atmosphere as well
as the water column. On the basis of project data presented in this report,
contaminant release to the water column might not depend on observable resuspension of solids (see Box 4-2 for an example). Nevertheless,
monitoring of turbidity can provide real-time quality-assurance information to the dredge operator and allow adjustments in the field to reduce
resuspension. Air monitoring can also identify potential exposures and
facilitate needed operational adjustments to protect nearby populations.
To quantify contaminant releases, however, upstream and downstream water-column contaminant fluxes should also be monitored. This
can be accomplished relatively quickly using immunoassay test kits or
traditional grab sampling with subsequent analyses. Passive sampling
devices or caged fish can also be placed at the site to indicate exposure
over extended periods. These techniques make it possible to quantify the
unintended contaminant loading to the water column, and this helps to
explain increases in downstream exposures that are observed between
the baseline and long-term monitoring and to distinguish short-term effects due to dredging from continuing long-term releases attributable to
uncontrolled sources.
Monitoring Human Health Effects During Dredging
During dredging and dredged-material handling, the surrounding
community and remediation workers might experience higher exposures
than before dredging. The increases could arise from chemical releases to
the overlying surface water and ambient air, from uptake by biota consumed by people, and from creation of residual contamination in areas
where people or edible biota come into contact with it. The surrounding
community might also experience non-health-related effects, such as accidents, noise, and residential or commercial disruption, which are potential ancillary consequences of dredging.
An evaluation of net risk reduction should begin with sufficient
datasets that permit comparison of exposure conditions before and dur-
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ing dredging operations. In some of the projects, increased exposure occurred during dredging in connection with the physical disruption of
contaminated sediment. Monitoring during dredging should be designed so that data are sufficient to quantify changes in exposure resulting from the dredging operation, specifically related to resuspension of
sediment, release of chemicals from sediment, and creation of residuals.
Changes in net risk resulting from transport, storage, treatment, and disposal of dredged sediments should be quantified, and this may include
collection of monitoring data during dredging and dredging-related operations.
Superfund remedial investigations often use concentrations of
chemicals in environmental media—such as fish, sediment, surface water, and air—as surrogates for human exposure and do not study human
subjects directly. Investigators need to monitor those media within the
boundaries of the three-dimensional space in which people have direct
or indirect contact. For example, people do not have direct contact with
deep sediment. Unless that sediment becomes exposed in the future as a
result of scouring, the dredging process itself, or some other process,
sediment samples collected for evaluating direct contact should not exceed the depth that a swimmer or wader might encounter. For indirect
exposure to stable sediments through the food chain, the relevant sediment sampling depth is limited to the biologically active zone. Sampling
at greater depths is needed to assess potential exposures where sediments are unstable. If sediment contamination has reached the terrestrial
environment through atmospheric release and deposition or sediment
deposition on floodplain soils, parallel monitoring and risk analyses
should be performed for terrestrial exposure media. Direct studies of
human exposure could help to quantify human exposure and risk but
would be more invasive and expensive and would not necessarily yield a
good measure of exposure reduction, given the difficulty in defining the
exposed population and segregating site-related exposures from other
exposures to chemicals of concern.
Given that dredging remediation by definition involves an aquatic
environment and that many of the most important sediment contaminants are bioaccumulative, the consumption of fish and other aquatic
organisms often contributes most to human health risk. However, until
dredging is completed and cleanup goals have been met, EPA and state
agencies with fisheries jurisdiction usually restrict fish consumption to
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protect human health. When members of the surrounding community
comply with those restrictions, exposures of concern during dredging
are limited to other pathways, such as releases to the atmosphere and
surface water. Box 5-3 highlights examples of attempts to evaluate human exposure during dredging.
Ideally, dredging operations occur over a relatively short period
that requires evaluation of acute and possibly subchronic risk, but not
chronic risk, from these exposure pathways. To address those risks, EPA
can establish acute and subchronic guidelines for air or other media for
comparison with monitoring results. For example, at the New Bedford
Harbor Superfund site, EPA selected air concentrations that if exceeded
during dredging would require a change in the dredging operation. Also
at that site, EPA detected increased hydrogen sulfide concentrations in a
dredged sediment handling facility and changed the operation to reduce
concentrations to safe levels.3 In such cases, monitoring results should be
made available in a time frame that allows site managers to manage risks
appropriately. The data should distinguish conditions upstream and
downstream of dredging or upwind and downwind of dredging so that
site managers can discern the effects of dredging relative to background
exposure conditions including natural disturbances.
Monitoring Ecologic Effects During Dredging
Current practice often omits biologic monitoring during remedy
implementation at sediment sites. Monitoring of bioaccumulation during
dredging is typically not able to inform the project manager or operator
in a timely fashion so that dredging protocols could be modified or additional protections implemented, owing to the length of time that most
organisms take to respond to environmental exposures. Other challenges
in using bioaccumulation and tissue concentration monitoring data are
in relating chemical concentration to ecologic relevance or adverse biologic effects and in the uncertainty of the relationship between exposure
(such as to site sediments or resuspended materials) and tissue concentrations in fish if they are able to move off site (these issues are described
in greater detail in the next section).
The presence of hydrogen sulfide is related to the anaerobic environment of
the sediments and not to chemical contaminants at the site.
3
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BOX 5-3 Monitoring of Conditions During Hot-Spot Dredging at the New
Bedford Harbor Superfund Site for Effects on Human Exposure
The New Bedford Harbor Superfund site has been the subject of extensive
efforts to understand the effects of harbor contamination on aquatic species and
people living near the harbor. In addition, EPA developed a plan to monitor the
effects of dredging a contaminated hot spot on water quality, air quality, and
bioaccumulation by benthic invertebrates. The hot-spot dredging occurred in
1994-1995. EPA (1997) compared results of monitoring conducted before, during,
and after hot-spot dredging and concluded that the dredging resulted in few if
any adverse effects on the marine ecosystem. EPA identified some air-quality
issues that were remedied with changes in operation or engineering controls.
Cullen et al. (1996) compared PCB concentrations in tomato samples collected downwind of the hot-spot dredging operation before and during dredging and concluded that the average PCB concentration during dredging was
about 6 times higher than the average PCB concentration before dredging.
Choi et al. (2006) reported PCB concentrations in umbilical-cord blood
samples among members of nearby communities that were collected before, during, and after hot-spot dredging. The authors reported that their results “support
modest, transient increases in cord serum PCB levels during dredging, with significant declines in serum PCB levels observed after dredging, particularly for
the more volatile PCBs and PCB-118.” They attributed the “significant declines,”
in part, to the hot-spot dredging (see figure below).
Figure shows covariate-adjusted smoothed plots of predicted ∑PCB (A), heavy PCB (B),
light PCB (C), and PCB-118 (D) levels vs. infant’s date of birth. Vertical lines denote the
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BOX 5-3 Continued
start and stop dates for dredging of contaminated New Bedford Harbor sediments. Plots
are adjusted for child’s sex, maternal age, birthplace, smoking during pregnancy, previous
lactation, household income, and diet (consumption of organ meat, red meat, local dairy,
and dark fish). Source: Choi et al. 2006.
EPA’s summary of tier 1 sediment remediation sites with pre-monitoring
and post-monitoring data (S. Ells, EPA, unpublished information, March 22,
2006) presented during the committee’s first meeting indicates that the baseline
average concentration of PCBs in surface sediment (no depth specified) at the
hot spot was 25,000 mg/kg and lists a post-remedial average concentration the
hot spot of 330 mg/kg. However, that information appears to be inconsistent
with another EPA presentation (Nelson 2006) during which W. Nelson reported
that the average concentration of PCBs in surface sediment (top 2 cm) of the upper harbor, of which the hot spot makes up about 5% of total area, did not
change significantly as a result of hot-spot dredging. It is possible that the average was unchanged because the hot spot represents a small fraction of the upper
harbor. However, if the hot spot represented the portion of the upper harbor
with the highest PCB concentration, one would expect its removal to cause some
decline in PCB concentration. Given that the PCB concentration apparently did
not decline, it is not clear how hot-spot dredging might have led to reduced PCB
concentrations in umbilical-cord serum after dredging.
These collective efforts show how exposure might change during the period of dredging, but it is premature to use them to judge effectiveness, because
EPA’s remediation is not yet complete. These studies illustrate the challenge of
linking dredging in a large harbor with human exposure.
Nevertheless, one of the main risks to ecologic receptors posed by
release and transport of contaminants during dredging is increased contaminant uptake and increased toxicity, and there are techniques for
monitoring those effects, even if they are not in wide use. Subject to the
timeliness limitations noted above for providing real-time feedback to
dredging operations, studies can be designed to assess contaminant bioaccumulation and toxicity during dredging by using caged or sessile organisms or using passive sampling devices such as SPMDs, as discussed
above (Chappie and Burton 2000; Adams et al. 2005; Crane et al. 2007).
The utility of caged-fish studies has been demonstrated, for example, in
the 1995 Grasse River non-time-critical removal action (BBL 1995). Mus-
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sels deployed in mesh bags have been used in the long-term monitoring
program at the New Bedford Harbor Superfund site to monitor trends in
PCB bioaccumulation and evaluate the impact of dredging operations
(Bergen et al. 2005). To quantify the spatial distribution of resuspended
materials, the organisms can be placed at various distances from the
sources of contamination. Comparisons with pre-dredging, reference, or
upstream conditions allow managers to determine whether uptake of
contaminants increases during dredging operations.
Complex exposure dynamics cannot be mimicked in the laboratory.
If standard test species are exposed in situ, exposures are more realistic.
In situ testing with caged organisms has been shown to be an effective
monitoring tool. Its primary advantages are the improved realism of exposure, the lack of sampling-induced artifacts, the ability to deploy and
assess within days, and the ability to partition exposures of key environmental compartments and exposure time frames. (However, these
techniques also have concerns regarding the modification of site conditions during exposure [Chappie and Burton 2000]). One can also link
exposure with effects in that multiple end points can be assessed, such as
tissue concentrations, growth, and reproductive status. Numerous studies have demonstrated the approach in studies of runoff, base flow, and
sediments (e.g., Ireland et al. 1996; Tucker and Burton 1999; Chappie and
Burton 2000; Greenberg et al. 2002; Burton et al. 2005; Crane et al. 2007).
Studies of marine systems have primarily used mussels (Salazar and Salazar 1997; Bergen et al. 2005), and there has been less testing of amphipods (DeWitt et al. 1999). Freshwater studies have used a wide variety of
organisms, such as fish, cladocerans, amphipods, midges, bivalves, mayflies, and oligochaetes (e.g., Chappie and Burton 2000). It is important to
consider the likely response time when selecting test organisms; organisms that equilibrate with their environment more quickly would be
more useful for evaluating releases during dredging. It is also advantageous, when possible, to use indigenous biota when conducting in situ
caged testing. Standard toxicity test organisms (such as fathead minnows) may have very different biologic responses than indigenous populations that may have acclimated or adapted to toxics in the watershed,
thus being less sensitive than the surrogate species. In that case, surrogate species may be useful for detecting adverse effects, but they will not
be a good indicator of effects to indigenous species.
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The use of natural resident populations collected during and immediately after dredging is an alternative to caging studies. For example,
resident spottail shiners collected during the 2005 dredging operations at
the Grasse River showed significant increases compared to sampling
conducted during several years prior to dredging and the year following
dredging (see Chapter 4, Figure 4-10 and associated text).
Post-dredging Verification Sampling
Verification sampling immediately after dredging allows site managers to determine whether cleanup levels or other short-term objectives
have been met. The ability of the remedy to achieve short-term cleanup
levels depends in part on how accurately the remedial investigation and
additional pre-remedial sampling have characterized the extent and distribution of contamination and on whether the dredging design based on
that characterization encompasses the bounds of contamination encountered by the dredger in the field. Dredging designs are based on interpolation of sediment core data, which are often sparse relative to the scale
of sites. Even at major sites, the density of pre-design samples is typically
less than one core per acre: for development of the Lower Fox River and
Green Bay Operable Units 2-5 remedial design, the density was about
one core per 1.6 acres (Shaw 2006). Depths of contamination in the wide
expanses between core locations are therefore subject to uncertainty, and
the dredging projects in this report provide evidence of that uncertainty
in the form of sites where significant undisturbed residuals remained
after dredging to design elevations. The probability of leaving consolidated sediments with elevated concentrations in place can be reduced by
conducting more intensive pre-design sampling before dredging, by
lowering the elevation of the dredge cut (that is, overdredging), or by
verification sampling after dredging followed by redredging as needed
(see examples in Box 5-4). Thus, there are tradeoffs between the volume
of material removed and the intensities of pre-design and verification
sampling. The greater the confidence in the methods used to develop the
dredge prism, including sampling and interpolation, the less overdredging and verification sampling may be needed to ensure protection. Those
tradeoffs have been considered explicitly in pre-design studies for Lower
Fox River Operable Units 2-5 (see Box 5-5).
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The dredging projects evaluated by the committee include numerous examples of sites where dredging generated substantial residual contamination. Verification sampling is needed to detect and quantify generated residuals. Where possible, the samples should be collected in the
form of cores long enough to penetrate and capture sediment underlying
the generated residual layer, rather than grab samples (Palermo 2006).
When cores cannot be obtained, grab samples should be taken. One
promising technology for obtaining grab samples, using a hydraulic
sampling device, is described in Box 5-6. It is also important, during collection and analysis of cores, to capture the unconsolidated “fluff” that
may be generated from dredging activity. With core samples, the thickness and texture of the generated residual layer can be observed and distinguished from underlying material to support planning of additional
work that may be needed to minimize risk, such as backfilling with
coarse-grained material.
BOX 5-4 Verification Sampling at Harbor Island, Washington
Dredging at Todd Shipyards (part of the Harbor Island Superfund site) relied on collection of shallow progress cores in each sediment management area
(SMA) as dredging was completed. Results were compared with cleanup levels
to determine whether additional dredging was needed. Dredging was sequenced
in such a way that an SMA that had been remediated was not affected by dredging in adjacent SMAs. Final verification samples were collected once all SMAs
had been dredged at a relatively low density because of demonstrated compliance with cleanup levels based on the progress cores.
The presence of extensive surface and subsurface debris in the areas to be
dredged at Lockheed Shipyard (also part of the Harbor Island Superfund site)
resulted in extensive residual contamination and undredged inventory at the
end of the first dredging season. Shallow cores were collected throughout the
dredged area after dredging to document the remaining contamination and distinguish between a light unconsolidated sediment layer and more consolidated
material. The latter material either had sloughed from the edge of the dredge cut
or could not be removed by the dredger because of debris that remained on the
site. On the basis of this sampling, it was decided to place a thin layer of clean
material over the dredged area until it could be redredged in the following season.
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BOX 5-5 Delineating the Dredge Prism in the Fox River, Wisconsin
Sediment remediation areas and volumes were delineated for the remedial
design of the lower Fox River, WI, between Operable Unit 2 and the mouth of
Green Bay with an advanced interpolation method called full-indicator kriging.
Full-indicator kriging provided a probability distribution of depth of contamination to the ROD cleanup level at each sediment location. For areas where dredging is the selected remedy, dredge-prism designs were developed at a range of
significance levels (defined as the probability of exceeding the cleanup level at a
given location) to inform risk management decisions. Those decisions involve
balancing the risk of leaving contaminated sediment behind (a false-negative, or
type 2, error) against the risk of unnecessarily dredging clean material (a falsepositive, or type 1, error). Additional protection against false negatives will be
provided in the remedy by post-dredging confirmation sampling. A significance
level of 0.5 was chosen because it provided a reasonable Type 1 error and a low
Type 2 error, reasonable accuracy, and the least bias in the dredge cut. This decision was made acknowledging the importance of minimizing Type 2 errors because remediation of clean sediments is cost that cannot be recovered. There was
agreement that a robust verification sampling program would be developed and
this would uncover significant Type 1 errors (that is, leaving behind sediments
that should be remediated) which can subsequently be dealt with. In practice,
more sediment will be removed than the selected significance level indicates
because additional deepening of the dredge prism occurs during dredging-plan
design and to account for contractor overdredging allowance (Anchor and
Limno-Tech 2006a,b,c).
In verification sampling, it is important that the spatial scale of
remedy evaluation be consistent with the site’s remedial objectives. If the
objectives require minimizing contaminant flux to the water column or
minimizing sitewide exposures to widely ranging fish species, it is appropriate to compare area-weighted average concentrations with remedial goals. Protecting sensitive receptors that have more limited ranges
would require verification that targets are achieved on a finer spatial
scale. It should also be emphasized that although effectiveness of implementation, which is an intermediate goal, can be evaluated with verification sampling, the ultimate goal is risk reduction through achievement of remedial objectives, which is evaluated with long-term
monitoring.
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BOX 5-6 Verification Sampling of Dredging Residuals at the Head of Hylebos
Site, Commencement Bay, Washington
Discrete sediment samples were collected on a daily basis immediately behind the operating dredge to provide immediate evaluation of the post-dredging
residual layer for the Head of Hylebos project. Nearly 1,000 discrete samples of
the residual layer were collected using a Marine Sampling Systems 0.3m2 Power
Grab (Power Grab) generating measurements of residual layer thickness and
sediment chemistry (24-hour chemistry turn around times). This program provided immediate feedback on the nature of the residual layer generated during
dredging, and allowed for ongoing adjustment of the dredging methods to further control the residual layer formation.
Unlike typical surface grab samplers, the Power Grab is a hydraulically actuated clamshell bucket that is capable of collecting 1-ft thick samples in many
sediment types ranging from soft fine-grained sediment to more dense and compact silts and sands (not hardpan or glacial till), as well as through some debris.
All of the sample contact surfaces on the Power Grab are stainless steel while the
frame of the sampler is aluminum. The Power Grab features include (1) wide
adjustable feet to control the depth of penetration to avoid over or under penetration of the sampler; (2) adjustable ballast (280-750 lb) to provide additional
reaction weight for sampling in stiff material; (3) a semi-circular cutting profile
to limit the disturbance of the sample; (4) and an enclosed bucket configuration
to protect the sample from scour while being raised through the water column.
These features allowed the sample team to consistently collect acceptable samples (without over- or under-penetration) in all sediment types found on the site.
Once the Power Grab sample was brought on board the sampling vessel,
the overlying bucket covers were removed and overlying water decanted. The
0.3-m2 sample footprint (roughly 1 3/4 ft by 1 3/4 ft) was sufficient to allow for
subsampling to measure the thickness and record the characteristics of the residual layer, the characteristics of the underlying more compact native sediment,
as well as the collection of sediment samples for chemical analysis. The Power
Grab performed well throughout the two seasons of dredge confirmation sampling without any notable complications or problems.
Source: Dalton, Olmsted & Fuglevand 2006.
Long-Term Monitoring after Dredging
Long-term monitoring is used to judge whether a remedy is reducing risk at the expected rate and when remedial objectives have been
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achieved. Superfund remedies at sediment sites are typically subject to
review at 5-year intervals when, following remediation, contamination
exists that could limit potential uses of the site (EPA 2001b). At dredging
sites this could occur for several reasons: residual contamination after
the completion of the remedial action, the recontamination potential associated with the dynamic nature of the aquatic environment, the fact
that some sources may be undetected and that controls of known sources
are not always implemented concurrently with the remedy (particularly
at the watershed level), and the additional time required by remedies to
achieve objectives when they rely in part on natural recovery processes
and must counter past bioaccumulation of contaminant in the food
chain. Ideally, reviews compare recovery at each 5-year interval with an
expected trend of exposure and risk reduction under the recommended
remedy as developed in the feasibility study. Remedy modification and
additional data collection as needed to fill in gaps in understanding,
would be triggered by a significant deviation from the expected trend.
Otherwise, monitoring continues until remedial objectives are achieved.
Although a rich set of data should already exist as a product of the
remedial investigation, it is important that a complete baseline dataset be
obtained before remedy implementation, observing the same pathways
and exposures as planned for long-term monitoring, to support clear and
definitive pre-remedial vs post-remedial comparisons and post-remedial
trend estimates. The importance of establishing a baseline, especially to
assess the effects of the remedy on fish, was emphasized by a previous
National Research Council panel (NRC 2001). Because long-term monitoring may continue for decades and trend estimates will be based on
comparisons with data collected in the early years of monitoring, it is
important that sampling and analytic methods selected for baseline and
long-term monitoring be consistent with the technologic state of the art.
It is also vital that the baseline data be collected before the commencement of remediation and encompass trends of sufficient duration for the
effects of the remedy to be distinguished from ambient trends leading up
to its implementation. To facilitate comparisons of data over a long period (decades), it is useful to store duplicate biologic samples (fish tissues, human tissues, blood) from the analyses (for example in a deep
freezer or liquid nitrogen). These samples can be analyzed in later years
to facilitate comparisons of analytical data.
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When a dredging remedy is implemented, surface sediment concentrations can be affected by a combination of sediment removal, backfilling with clean material, and natural recovery processes. At larger
sites, where remediation may proceed over a period of years, there is
value in determining the relative importance of each of those processes
in reducing surface concentrations. If burial under clean watershed
sediment transported by riverine processes strongly reinforces dilution
of exposures, this may create opportunities to adapt the remedy to reduce cost without compromising effectiveness. For example, if burial by
clean sediment is sufficiently rapid and uniform, it may be possible to
achieve risk-reduction goals while tolerating higher generated residual
concentrations or to reduce the thickness of post-dredging backfill (see
Box 5-7). To be able to measure residuals, backfill, and long-term sedimentation as separate layers, baseline and long-term monitoring of surface sediment should include at least a subset of finely sectioned cores,
which should be analyzed for geotechnical characteristics, including
grain size, bulk density, and contaminant concentrations.
It is important to stress that backfilling and burial by natural sediment processes are not necessarily equivalent in protection even if they
cover contamination to equal thicknesses. Sands, which are typically
used as backfill material, may be more stable in the face of erosion during high-flow events than natural sediments but their lower sorptive capacity may provide less effective attenuation of contaminant exposure
than burial by natural sediment, especially when pore water is the pathway of potential exposure.4 Monitoring of surface concentrations should
account for trends in bioavailability by normalizing contaminant concentrations to sorbent material, such as organic carbon, in addition to measuring trends in bulk surface-sediment contaminant concentrations.
Long-term Monitoring of Human Health Effects
Results of the baseline human health risk assessment indicate
which exposure pathways and chemicals warrant action, and this knowledge is vital for making effective remedial decisions. As noted above,
the consumption of fish and other aquatic organisms often contributes
4
It also is possible to specify backfill material with organic carbon.
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BOX 5-7 Dredging and Later Sedimentation at Manistique Harbor, Michigan
After dredging ended at Manistique Harbor, 3-7 ft of sediment were deposited from 1996 to 2005 (Weston 2005). With the deposition of the new surface
sediment on post-dredging residuals, it was possible to meet revised dredging
cleanup levels. The original cleanup level had been removal of all sediment containing PCBs at greater than 10 mg/kg anywhere in the sediment column, and it
proved difficult to achieve. The cleanup level was revised to an average concentration of 10 mg/kg throughout the sediment column, with 95% removal of PCB
mass also required (see Appendix C).
most to human health risk. Ideally, the population consuming aquatic
biota would have been studied in the baseline human health risk assessment, including quantification of variability in fish consumption
rates among members of the population to ensure that the most highly
exposed members of the population are evaluated in the risk assessment.
During monitoring, any important changes in consumption patterns
should be accounted for in the monitoring plans. However, other pathways might be important, such as consumption of waterfowl, dermal
contact with and ingestion of sediment, inhalation of volatilized contaminants, and ingestion of surface water during swimming or other recreational activities or through use as drinking water. If sediment contamination reaches the floodplain, people could be exposed through
consumption of game species, wild edible plants, and agricultural products from the floodplain and through dermal contact with and ingestion
of surface soil. Box 5-8 summarizes factors that one should consider in
designing the aquatic sampling programs that are most often used to
quantify human exposures.
EPA’s goal at contaminated sediment sites is to protect human
health, given that people could be exposed to sediment and other contaminated media over long periods. Consequently, cleanup levels, if established to protect human health, usually represent long-term average
concentrations that people can be exposed to without expectation of
harm over long periods. Therefore, they should not be treated as absolute exposure limits that are not to be exceeded at any time during longterm monitoring. Cleanup levels and monitoring programs should be
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BOX 5-8 Collecting Aquatic Samples for Monitoring Human Exposure
Fish and Shellfish
• Sample the species commonly eaten by the local population and be sure
to include species known to accumulate high concentrations of chemicals of concern.
• Catch the size range of fish harvested by the local population, being sure
to include the larger fish usually harvested because larger (older) fish in a population are generally the most contaminated with chemicals that bioaccumulate
(such as PCBs, dioxins, and methylmercury).
• Avoid sampling finfish species during their spawning period, because
tissue concentrations of some chemicals (for example, such lipophilic chemicals
as PCBs and dioxin but not methylmercury) may decrease during this time and
because the spawning period is generally outside the legal harvest period.
• Match assumed or known consumption patterns to sampled species.
Fish-creel data (from data gathered by surveying anglers) from state fisheries
departments constitute one justifiable basis for estimating types and amounts of
fish consumed from a given body of water. It is important to account for the fractions that various trophic levels contribute to a fish consumer’s diet.
• Composite samples of fish parts consumed by the local population. People might eat skin-on fillets, skin-off fillets, or whole gutted fish (for example, in
soups). Skin-off fillets will have the highest mercury concentrations, whereas
whole-body fish samples will have the highest PCB and dioxin concentrations.
PAHs do not tend to accumulate in finfish that metabolize them. Composites
improve the chance of detecting chemicals and thus reduce the number of samples without detectable concentrations in the resulting dataset and the need to
determine how they will be factored into arithmetic averaging.
• Use a probabilistic sampling design, randomly selecting sampling locations to address spatial variability and to ensure that sufficient samples are collected to distinguish the site from reference areas. This approach allows statistically valid inferences to be drawn on an area as a whole. Ideally, samples would
be collected over a geographic area that represents the average exposure of those
who eat fish from the body of water. If there are smaller areas where people are
known to concentrate fishing, these areas should be intensively sampled.
• Collect both weight and length data to control for the potential influence
of fish nutritional state on chemical concentration, such as by normalizing fish
concentrations to a standard body condition.
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BOX 5-8 Continued
Sediment
• Collect from accessible locations where people are likely to fish, swim, or
engage in other activities (sediment samples in deep water, for example, may be
relevant to the food-chain exposure pathway but not the direct-contact pathway).
• To evaluate direct-contact exposures, collect sediment at depths that correspond to the depth to which a swimmer or wader might sink.
• To evaluate indirect food-chain exposures, collect sediment from the biologically active zone.
Surface Water
• Collect from accessible locations where people are likely to fish, swim, or
engage in other activities.
• Collect from areas used as a drinking water source.
• Measure total chemical concentrations if people ingest the water. Dissolved-phase concentrations are more useful for some evaluations of dermal
exposure.
Source: Adapted from EPA 2000c.
defined in the context of areas over which people average their exposure.
For example, if the pathway of concern is direct contact with sediment,
concentrations in sediment that are routinely beneath 10 ft of water are
of less concern than concentrations in shallow, accessible waters at the
shoreline. Cleanup levels should be compared with uncertainty bounds
on average exposures, such as the 95% upper confidence limits of the
mean concentration in each human exposure area, rather than the maximum concentration detected in each exposure area.5
The conceptual site model and feasibility study results should be
used to set expectations for the rate of risk reduction. To ensure that unacceptable risks do not occur, site managers can track concentrations
One caveat to the safety of long term averages and the relative unimportance
of short term exceedances is that exposure can occur during a vulnerable period,
such as fetal development or infancy.
5
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monitored over time to estimate expected cumulative cancer risk and
noncancer hazards. If monitoring during or after dredging indicates that
cleanup levels will not be met in the long run, site managers can use
adaptive management to change this trend.
Long-term Monitoring of Ecologic Effects
A primary goal of long-term monitoring is to test the hypothesis
that dredging has reduced injury to ecologic resources. Many of the
techniques used to assess potential short-term adverse effects of dredging on ecologic resources, described above, are also appropriate for assessing long-term effectiveness of sediment removal. For example, benthic toxicity testing has been successfully used as part of long-term
monitoring at the former Ketchikan Pulp Company site in Alaska. In
Waukegan Harbor, laboratory sediment toxicity assays conducted four
years after dredging showed reduced toxicity compared to pre-dredging
assessments, but nevertheless, toxicity still persisted (Ingersoll et al. 1996;
EPA 1999; Kemble et al. 2000). Followup studies in the Black River
showed that surficial sediments had reduced toxicity but PAHs were still
causing toxicity in caged organisms (Burton and Rowland 1998). Contaminant uptake and toxicity can also be quantified by using in situ approaches with caged organisms or using passive sampling devices such
as SPMDs, as discussed above. In addition to placing cages or SPMDs at
different locations to quantify spatial distribution of contaminant concentrations or effects, observations can take place over time to determine
temporal changes in bioavailability and sediment toxicity. SPMDs were
used in Manistique Harbor 4 years after dredging operations ceased. The
SPMDs accumulated PCBs to detectable levels while PCBs were not detected in caged fish or surface water samples at the site (Weston 2005).
Long-term monitoring of resident populations of fish and invertebrates can also reveal changes in contaminant concentrations and ecologic effects resulting from removal of contaminated sediment and natural
processes. For example, the incidences of lesions and tumors in brown
bullheads in the Black River showed initial increases and then marked
reductions following dredging at the site (Baumann 2000) compared
with pre-dredging conditions. Tissue data from fish whose habitat is limited to the remediation site are valuable indicators because they integrate
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exposures over the remediation area. Several of the dredging projects
(such as Waukegan Harbor, Grasse River, Black River, the Puget Sound
Naval Shipyard, and GM Massena) that were evaluated by the committee monitored fish tissue concentrations of contaminants to evaluate the
effectiveness of sediment removal. Distinguishing the effects of remediation from background trends on the basis of fish tissue data often proves
problematic for the decision-making process because of the scarcity and
variability of fish tissue data. In addition, the difficulty in quantifying the
movements of fish in and out of a project area can make linkages between exposure and effects problematic. It may be impossible to determine how much time a fish has been exposed to study-site sediments
compared with offsite sediments (which may also be contaminated). Bioaccumulation modeling approaches (e.g., Linkov et al. 2002) that include
spatial and temporal characteristics of exposures based on a knowledge
of the organism’s life history patterns can be useful in addressing that
issue. Furthermore, pre-dredging data are often limited to a short time
and very few fish (such as at Waukegan Harbor), so it might be impossible to make statistically valid comparisons of trends (as discussed in the
next section). In these situations, caged-fish studies can maximize exposure to test-site sediments and thereby reduce uncertainty. Fish may be
the receptors of primary interest for both ecosystem and human health
risk, but monitoring them supports effective decision-making only if sufficient samples are collected and their patterns of exposure are known.
For determinations of ecosystem and human health risk, it is sometimes more effective to monitor tissue concentrations in benthic invertebrate organisms that reside at the test site (Adams et al. 2005; Burton et
al. 2005; Solomon et al. 1997) because these organisms tend to be sessile
or relatively immobile (for example, mussels). The organisms are exposed to the contaminated sediments through direct contact and are a
food source for fish, birds, and mammals; therefore, food-web transfer
and risk can be (and have been) modeled. The uncertainty of exposure is
largely removed, and organisms are easier and less expensive to collect
than fish and provide a convenient surrogate, as long as any assumed
bioaccumulation link can be verified with site data. In addition, passive
sampling devices that are biomimetic have recently been successfully
used (see above discussion). The adsorption of organic and metal contaminants on these devices has been shown to be similar to that of tissue
concentrations in indigenous organisms, so they can be used as a surro-
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gate for fish (Arthur and Pawliszyn 1990; Huckins et al. 1990; Zhang et
al. 1998; Wells and Lanno 2001; Lanno et al. 2004, 2005). Bioaccumulation
and toxicity studies can also be conducted in the laboratory with sediment and water collected from field sites after dredging operations.
However, it is critical that laboratory studies consider abiotic factors that
may influence contaminant bioavailability and degradation—such as
ultraviolet light, suspended solids and colloids, and organic carbon—
and the effect that removing sediments from the environment will have
on bioavailability.
Monitoring the structure and composition of benthic macroinvertebrate communities is a common approach to the assessment of effects of
sediment contaminants and can be used when the benthic community is
an important component of the conceptual site model of increased site
risks. Such community characteristics as total abundance, species diversity, richness, and abundance of sensitive species can be compared with
pre-dredging data and, when possible, with nearby reference sites.
Again, it is critical that similar methods be used to collect and process
pre-dredging and post-dredging samples. One of the greatest challenges
associated with long-term monitoring of benthic communities is to separate effects of dredging from changes due to other environmental factors.
The condition of benthic communities is generally expected to improve
after the removal of contaminated sediment, as may be predicted by the
conceptual site model. The failure of benthic communities to recover after dredging could be a result of residual sediment contaminants, lack of
colonizing organisms, conversion to an inhospitable or unsuitable habitat, or the presence of other stressors (Kelaher et al. 2003). Monitoring
approaches using in situ cages that contain natural benthic communities
offer an opportunity to demonstrate causal relationships between stressors and ecologically relevant responses. Demonstrating changes in the
tolerance of populations or communities may also provide evidence of
effectiveness of dredging in situations where traditional community
metrics (such as abundance and species richness) do not show recovery.
For example, increased tolerance to metals is often observed at metalcontaminated sites (Weis and Weis 1989; Clements 1999), so the loss of
tolerance in a population or community after dredging is evidence that
remediation was successful (Levinton et al. 2003). These experiments are
a practical alternative to single-species toxicity tests and address the sta-
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tistical problems associated with field biomonitoring studies (Clark and
Clements 2006).
The rate of recovery of benthic communities after dredging will be
determined by both biotic and abiotic factors (Yount and Neimi 1990).
The rate of recovery will be influenced not only by the adverse effects of
large-scale substrate disturbance and the presence of residual contaminants, but also by proximity to reference areas and the availability of
colonizing individuals. Ecosystems that have a direct connection to clean
reference sites will probably recover faster than closed systems that have
relatively little exchange. For example, the relatively fast recovery of
stream ecosystems after remediation has been attributed to rapid colonization by organisms from upstream reference sites (Clements and Newman 2002).
DATA SUFFICIENCY AND STATISTICAL DESIGN
Monitoring datasets should be rich enough to support testing of the
hypothesis that dredging is effective in meeting its remedial goals and
objectives. That requires that sampling targets the important exposure
pathways and be designed to capture temporal and spatial variability
and that sample sizes be sufficient for robust hypothesis-testing and statistical modeling of dredging effectiveness goals.
Standard statistical tests are often formulated as a null hypothesis
representing no effect or no change vs an alternative hypothesis representing an effect or change. When evaluating dredging effectiveness on
the basis of pre- and post-dredging data, the null hypothesis represents
no change due to dredging; the alternative hypothesis is that there was a
change in environmental conditions because of dredging. Established
formulas exist (EPA 2000d) for sample size determinations based on that
traditional approach. With the required estimate of outcome variability
(as can be obtained, for example, in a pilot study) and specification of the
minimal effect size that should be detected, sample size determinations
are based on optimizing the two types of statistical errors that can result.
The probability of type I errors (incorrectly claiming dredging to be effective when it was not) is fixed to be small, as is commonly done by setting the probability at 0.05. The probability of type II errors (failing to
claim that dredging was effective when it really was) is minimized,
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maximizing statistical power (Mason et al. 1989). The approach is thus
conservative; the burden of proof is on the monitoring data to provide
enough data points to support dredging effectiveness with a high degree
of confidence. One-sided alternative hypotheses can also be considered
to test whether an effect or change was in a specific direction, such as a
significant reduction in site conditions, and can consider varying minimal effect sizes, such as a reduction in site conditions of at least 90%
from pre-dredging or related background values. The latter approach
can be compared with hypothesis-testing techniques based on the bioequivalence paradigm (McDonald et al. 2003).
Sample size determination is crucial, but other components of statistical experimental design should not be overlooked in developing
monitoring plans to evaluate dredging effectiveness. A clear scientific
definition of dredging effectiveness is needed so that appropriate statistical hypotheses can be formulated. Hypotheses to be tested should be
based on and fully informed by the conceptual site model of exposures
and risks, as developed in the remedial investigation and baseline risk
assessment. Outcome variables need to be established, and their spatial
and temporal support (where, when, and how much) should be determined; all this should be consistent with and inform the statistical hypotheses. Careful determination and measurement of potential sources
of variation that may affect outcome variables are also important. Two
sources of variation that deserve further focus are temporal and spatial
variation in dredging effectiveness and their influence on monitoring
and followup statistical analysis.
It has been well established in this chapter and in the dredging projects reviewed in the previous chapter that characterizing environmental
conditions with monitoring before and after dredging is an important
design consideration for evaluating dredging effectiveness. Less established are guidelines for determining when temporal characterizations
should be assessed and whether assessment should follow a crosssectional approach of one time before and one after dredging or be longitudinal and use multiple monitoring times before and after dredging.
With just two time monitoring points before and after dredging (with
multiple samples taken at each of these time points), one can determine
whether a significant increase or decrease occurred between the two time
points. When dredging (or another remedial action) takes place between
these points, it is often assumed that the change results from the reme-
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diation, however, it is essential to consider trends that would occur regardless of dredging (for example, natural decreasing or increasing
trends in contaminant concentrations in sediment or fish). The value of
monitoring at several time points before and after dredging is that any
trends not due to dredging can be determined and the effect of dredging
more clearly established. Examples of fish tissue analyses that would
have benefited from more complete time trend data are presented in
Chapter 4 Boxes 4-1 and 4-6 on the Grasse River and Waukegan Harbor,
respectively.
Spatial or geographic variability can be an important component of
overall variability to consider in designing monitoring plans to evaluate
dredging effectiveness. There could be several reasons for spatial variation in dredging effectiveness at a site. There could be naturally occurring variations in environmental factors, such as water flow, wind patterns, and sediment texture. There could be spatial variations in site
conditions that affect the ability to dredge or dredge effectively, such as
the presence of bedrock, harbor infrastructure, or debris. It is not only
important to collect location information with monitoring data but to
statistically inform the monitoring plan to determine appropriate locations of monitoring samples. For example, in analysis of the Grasse River
project (see Chapter 4), locations of sediment samples were too far apart
to identify spatial variation in surficial PCBs. Analyses of the Lavaca Bay
project (see Chapter 4) suggested that subarea variations in surface mercury were of interest, but samples were too small to test this hypothesis
statistically while also considering temporal variation after dredge
passes. The subfield of statistics known as geostatistics (for textbook
treatments, see Cressie 1991; Goovaerts 1997; Diggle and Ribeiro 2007)
deals with the design and analysis of spatially referenced data that
commonly arise in monitoring of dredging applications and should inform monitoring plans and data analysis.
Even when spatial variation in dredging effectiveness is not of primary interest, the data collected through monitoring may very well exhibit spatial dependence, that is, measurements of samples taken closer
together are more similar than those of samples taken farther apart.
Overlooking that property can result in hypothesis tests and statisticalmodel inference with biased levels of significance (Cressie 1991). Obtaining and including sample coordinates in monitoring databases will allow
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followup statistical analyses to include possible spatial dependence and
make the appropriate adjustments when necessary.
In the dredging projects reviewed, the committee found that the
quantity and quality of available and accessible monitoring data varied
considerably. Followup statistical analyses of monitoring data often were
nonexistent or consisted of simple summaries and graphs lacking any
formal notion of statistical uncertainty; this created a critical gap between the large expenditures devoted to monitoring and the ability to
provide scientifically defensible claims of dredging effectiveness based
on monitoring data. It is imperative that rigorous statistical analysis of
monitoring data be performed so that assessments of dredging effectiveness reflect the inherent uncertainties involved.
APPROACHES TO IMPROVING MONITORING
The dredging projects reviewed by the committee revealed limitations in the ability to make real-time adjustments in dredging operations
to minimize contaminant releases; to connect remedial actions with their
effects in space and time on exposure pathways, receptors, and ecosystems; to base monitoring on adequate conceptual site models of chemical
fate and transport and of human health and ecologic risk; and to understand the roles of multiple processes in determining effectiveness of
dredging. This section proposes approaches with promise to overcome
those limitations. Each will require method development and evaluation
before becoming part of the standard monitoring tool kit, and some may
be appropriate only in particular cases. The methods can contribute to a
weight-of-evidence basis of decision-making (Wenning et al. 2005) to
reduce uncertainty in evaluation of risk reduction at specific sites. The
topic of innovative monitoring methods is also reviewed by Viollier et al.
(2003) and Apitz et al. (2005).
Approaches with potential to improve site investigation and operational and post-remedial monitoring include the following:
• Measure sediment, pore water, and surface-water concentrations
rapidly and accurately.
• Monitor real-time contaminant releases during dredging.
• Measure the bioavailable fraction of contaminants in the field.
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• Closely link exposure data (that is, chemical data) with biologic
effect.
• Understand and model biologic uptake.
• Understand and model ecosystem response and recovery.
• Understand and model reduction in human exposure.
• Adequately and quickly identify generation, production, transport, and deposition of sediment residuals.
• Understand and model processes responsible for recovery after
dredging.
• Quantitatively account for uncertainty in predictions of risk reduction and in later monitoring.
Measure Sediment Pore Water Concentrations
Rapidly and Accurately
A growing body of evidence suggests that sediment pore water
concentrations are strong indicators of the effects of sedimentcontaminant concentrations on benthic organisms (Adams et al. 1985; Di
Toro et al. 1991; Jager et al. 2000; Kraaij et al. 2003; Wenning et al. 2005;
Lu et al. 2006). Sediment pore water concentration is directly related to
the amount of bioavailable contaminant and uptake by benthic organisms (McLeod et al. 2007). Current methods to measure sediment pore
water involve the equilibration of sediment samples in the laboratory
and extraction of equilibrated water or the use of biomimetic assays, as
discussed above. Rapid techniques for measuring sediment pore water
would provide more useful and timely information on the status of recovery and resulting reduction in risk to humans and ecosystems.
Method development and pilot testing are needed to determine how reliably the available techniques can be adapted for laboratory and field
conditions.
Monitor Real-Time Contaminant Releases in the
Field During Dredging
Cost-effective methods are needed for real-time monitoring of contaminant releases during dredging. At present, turbidity commonly is
used as a surrogate for the release of persistent organic contaminants
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(such as PCBs) because the measurement is robust and quick and may be
automated. It is often assumed that turbidity release is proportional to
contaminant release. However, that may not be the case, as was shown
in several of the Chapter 4 dredging project evaluations, because contaminant fractionation between the aqueous phase and sediment particles can result in releases of aqueous-phase (or colloid-associated)
chemical contaminants, depending on the size and chemistry of the solid
phase. Furthermore, if dredging exposed nonaqueous phases, such as
liquid tar or hydraulic oil, contaminant release from such phases to overlying water would not be related to the release of solids. In principle,
some of the latest methods to measure contaminants in sediment pore
water—such as ELISA, PEDs, and SPME—could be applied to monitor
the release of contaminants to the aqueous phase if their detection limits
prove adequate. The newer methods in conjunction with turbidity, to the
extent that they are correlated, may facilitate more reliable and faster
contaminant monitoring during dredging.
Measure the Bioavailable Fraction of Contaminants in the Field
A variety of potential methods are candidates for measuring the
bioavailable fraction of organic and inorganic contaminants in sediments. This information, for both baseline and post-remedial sampling,
would complement chemical data to provide a more complete picture of
changes in exposure due to the remedy. The National Research Council
report on the bioavailabilty of contaminants in soils and sediments has a
long and detailed chapter devoted to this topic (NRC 2003). However,
most of the methods are not compound-specific or require detailed instrumental methods of analysis. Field methods that are compoundspecific are desired. One promising approach is immunoassay techniques for assessment of the bioavailable contaminant concentration.
Contaminant-specific ELISAs (Johnson and Van Emon 1996) may be
rapid, useful tools for measuring available contaminants. The immunoassay uses the selectivity and sensitivity of antibody recognition coupled
to an enzymatic reaction to rapidly determine chemical (such as PCB)
concentrations in a variety of media, including wet sediment extracts
and pore water. Ideally, the whole procedure, from extraction to colori-
220
metric detection, could be carried out in the field with a test kit and portable equipment [Ta 2001].
Understand and Model Biologic Uptake
Site models that are used to support site investigations, remedy selection, and remedial monitoring should include a model of contaminant
uptake by the affected biota. Biodynamic models describe the uptake of
contaminants as a mass balance of uptake from water; uptake from food
particles, including sediment; and loss rates. Such models would help to
explain the relationship between level of sediment cleanup and concentration of contaminants in organisms. The typical bioenergetics-based
toxicokinetic model (for example see Norstrom et al. 1976) assumes that
uptake by each route is independent and additive. The model has been
used to determine the uptake of contaminants by different routes with
experimentally determined model parameter values (Boese et al. 1990;
Weston et al. 2000; Lu et al. 2004). Luoma and Rainbow (2005) recently
proposed biodynamics as a unifying concept in metal bioaccumulation,
and similar formulations have been used in PCB food-web models (e.g.,
Connolly and Thomann 1992). It is proposed that a biodynamic model
that integrates sediment, water, and organism data from field projects
with the rapid assessment techniques described above be used to predict
contaminant concentrations in several species of interest. For example,
McLeod et al. (2007) showed that this biodynamic model successfully
predicted PCB body burdens in the clam Macoma balthica exposed to untreated and activated-carbon-amended Hunters Point sediment in laboratory experiments (see Box 5-9).
Understand and Model Ecosystem Response and Recovery
We lack rigorous modeling approaches to predict the ecologic
characteristics of recovery after sediment cleanup. It is possible that an
explanatory approach to ecologic recovery could build on the biodynamic modeling described above. If so, it could be incorporated into the
conceptual site model and used to support the site investigation, remedy
selection, and post-remedial monitoring. That belief is founded on the
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BOX 5-9 Biodynamic Modeling to Predict Organism
PCB Concentrations
If an organism is considered a single compartment for contaminant uptake, the following biodynamic equation describes its accumulation of a toxic
contaminant (McLeod et al. 2007):
dCorganism
dt
= FR ⋅ AE aq ⋅ Caq + IR ⋅ AE sed ⋅ Csed − k e ⋅ Corganism
where, Corganism is the contaminant concentration in soft tissue (µg/g dry), FR is
the water filtration rate (L of water per g dry per day), AEaq is the contaminant
absorption efficiency from water, Caq is the aqueous contaminant concentration
(µg/L), IR is the sediment-particle ingestion rate (g of sediment per g dry per
day), AEsed is the contaminant absorption efficiency from sediment, Csed is the
sediment contaminant concentration (µg/g dry), and ke is the proportional rate
constant of loss (per day). Model parameters include organism and filtration and
ingestion rates estimated from the literature. Sediment and aqueous contaminant
concentrations would be measured in situ, and laboratory experiments would
determine absorption-efficiency values and loss rates for the model organisms.
The advantage of this approach is that once the organism parameters values are
obtained, the conceptual model is transferable to other locations.
fact that contaminants in sediments simplify community structure by
eliminating some species but not others. Therefore, recovery should involve return of the contaminant-sensitive species to the community. In
addition, benthic communities in estuaries are dynamic in space and
time (Nichols and Thompson, 1985), so traditional ecologic observations
should be frequent and detailed to resolve community recovery. Biodynamic modeling based on functional ecology may allow prediction of the
species most sensitive to a contaminant, and this predictive capability
will help biologists to identify which species from the available recruitment pool are likely to recolonize a site when the contaminant is removed or bioavailability is reduced. Recolonization of the contaminantsensitive species in a recovering habitat reflects the success of remediation. A hypothesis to be tested is that recolonization predictions can be
built from basic information on taxon-specific functional ecology and
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biodynamics and contaminant metabolism, combined with data on species availability for community recruitment.
Understand and Model Reduction in Human Exposure
Monitoring programs should include measurement of surface
sediment, surface water, edible aquatic species, and other environmental
media found in the baseline risk assessment to present unacceptable
human health risks through either direct or indirect exposure. Monitoring determines whether exposure concentrations have declined as predicted. In addition, some systematic studies of the U.S. population, such
as the National Health and Nutrition Examination Survey (NHANES),
include biomonitoring data on some of the contaminants commonly detected at Superfund megasites. At the New Bedford Harbor Superfund
site, members of the surrounding community have been studied, including collection of umbilical-cord serum and breast-milk samples, as part
of an epidemiologic study of PCB effects on young children. Such human
biomonitoring studies can be expensive and invasive and are not entirely
without risk to those being monitored. Therefore, the committee does not
recommend implementing human biomonitoring sampling for all dredging projects. However, if relevant human biomonitoring data exist, they
can be reviewed for evidence that dredging resulted in reduced human
exposure and risk. Noninvasive biomonitoring might also improve future assessments of human exposure. For example, Fitzgerald et al.
(2005) reported a significant correlation between a noninvasive test of
enzyme activity related to PCBs and serum concentrations of PCBs in
members of a Mohawk tribe living near the General Motors-Central
Foundry Division Superfund site along the St. Lawrence River. Serum
PCB concentrations in this population had previously been correlated
with consumption of fish from the river (Fitzgerald et al. 1996, 1999,
2004).
Adequately and Quickly Identify Generation, Production, Transport,
and Deposition of Sediment Residuals
The purpose of sediment verification sampling is to ascertain
whether additional dredging passes, backfilling, or other remedial fol-
223
lowup is needed to meet risk-based cleanup levels. Downtime for dredging equipment and operators is expensive, but verification sampling and
laboratory analysis can be slow and laborious and require dredgers to
move on to other locations and return when results are available and
have been reviewed. Methods include grab sampling, coring, and visual
inspection by diver or with an underwater camera.
Operator response to verification sampling is limited by best
achievable laboratory- turnaround times and would be improved by development of more reliable field methods of analysis and greater use of
mobile laboratories. In combination with the most rapid methods of
analysis, the development of correlations between target chemical concentrations and sediment physical properties, as may be reflected in
sediment layering and other geomorphologic features, has the potential
to streamline sampling and analysis and to provide cost efficiencies and
much more rapid feedback to operators (Dow 2006).
Understand and Model Processes Responsible for
Recovery After Dredging
As discussed above, a dredging remedy can affect surface sediment
concentrations through a combination of sediment removal, backfilling,
and enhancement of natural recovery processes by creating areas of
preferential settling and deposition. Backfilling in particular was a component of the remedy at many of the sites considered in Chapter 4, but
its risk-reduction efficacy is uncertain and probably depends on the nature and thickness of the backfill material. Although backfilling provides
a separation layer between the water column and the contaminant, the
effective attenuation of exposure by backfill material may be minimal if
it is in a thin layer and has low adsorptive capacity.
To understand the long-term effects of dredging remedies on risk,
those issues should be evaluated as part of the monitoring program. Because undredged residuals, generated residuals, and backfill material
would be expected to differ in grain size and bulk density, these layers
should be delineated, after dredging, through physical and chemical
analysis of finely segmented cores. A time series of similar followup coring data, as part of the long-term monitoring program, would suffice to
distinguish the effects of post-dredging burial from those of backfilling.
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To estimate the combined effects of backfilling and burial on bioavailability, organic-chemical concentrations in surface layers should be organic-carbon normalized, and acid-volatile sulfide analyses of metals
should be conducted. Emerging field methods of pore water and bioavailability analysis should also be applied as they become more reliable
and widely available.
Account for Uncertainty Quantitatively in Predictions of Risk
Reduction and Later Monitoring
All risk assessments have inherent uncertainty in fate and transport
modeling and quantification of exposure and toxicity. The goal of monitoring should be to measure a given level of net risk reduction with a
reasonable degree of confidence. By acknowledging uncertainty, one is
better equipped to design an effective monitoring program and to answer questions from affected communities. For example, quantification
of uncertainty may enable site managers to inform community members
that the vapor-phase concentration of a chemical that will be released
during dredging operations and reach the nearest neighborhood, on the
basis of the best available modeling, is well below levels of concern at a
specific high level of confidence.
CONCLUSIONS
The committee draws the following conclusions concerning monitoring of dredging effectiveness in reducing risk:
• Monitoring is the only way to evaluate the success of a remedy
in reducing risk and is therefore an essential part of the remedy.
• Trends that occur at these sites are subject to biologic, chemical,
and physical processes that often operate on long time scales. The trends
and processes may be best described and understood with long-term
modeling and monitoring in pre-remedial and post-remedial time
frames.
• In the absence of sufficient baseline data, it is impossible to
evaluate effectiveness. Where pre-dredging conditions are not static, a
225
pre-remedial time-trend analysis is needed to judge remedial effectiveness.
• In most cases reviewed by the committee, monitoring has not
been adequately designed or implemented. Specifically,
o The design of the monitoring has often not been linked
sufficiently to the conceptual site model.
o Tools developed for the remedial investigation, including numerical models and baseline risk assessments, are
often neglected in formulating monitoring plans.
o Baseline datasets have not always been consistent with
long-term monitoring data.
o Contaminant exposure and effects have not always been
adequately linked in time and space.
o In many cases, the quality and quantity of monitoring
have been insufficient to support rigorous statistical
analyses.
• Some of the currently used monitoring techniques have proved
useful in determination of short-term and long-term effects of remediation. These include:
o Monitoring during dredging, such as measurement of
mass flux through upstream and downstream chemical
monitoring and biologic monitoring, including cagedfish studies.
o Long-term monitoring of fish tissue, where appropriate,
and other pathways that contribute substantially to human health risks.
o Long-term monitoring of affected benthic communities,
including tissue concentrations and health of benthic
communities.
o Laboratory toxicity testing using benthic organisms in
sediment to monitor long-term changes following
dredging.
• If fish are exposed to offsite conditions, there is uncertainty as to
their exposure and the relationship to risk. In those cases, benthic organisms may be better indicators of exposure, provided that their use is consistent with the conceptual site model of exposure pathways. If biologic
testing is not possible, passive sampling biomimetic devices provide indications of contaminant exposures.
226
RECOMMENDATIONS
The committee offers the following recommendations for improving monitoring of dredging effectiveness:
• EPA should ensure that monitoring is conducted at all contaminated sediment megasites to evaluate remedy effectiveness. That will
require a commitment of resources commensurate with the scale and
complexity of the site.
• Monitoring plans should focus on elements required to judge effectiveness and inform management decisions for the site. Care should
be taken to select the correct indicators of ecologic or human risk carefully. All aspects of monitoring—including planning, evaluation, and
adaptive management based on monitoring findings—should be closely
linked to the to conceptual site model so that the hypotheses and assumptions that led to the selected remedy can be tested and refined.
• The breadth and richness of monitoring datasets should be sufficient to support the testing and full evaluation of effectiveness goals. Statistical expertise should be included to inform well-designed monitoring
programs, guide database development, and perform rigorous statistical
analysis of monitoring data aimed at evaluating effectiveness.
• EPA should ensure that monitoring information on all Superfund megasites is systematically collected, organized, analyzed to assess
the effectiveness of remediation, and made available to the public in
such a form that effectiveness evaluations can be independently verified.
• Numerical models that are used in the remedial investigation
and feasibility study to design the remedy should be revisited during the
remediation phase to help in the evaluation of the effectiveness of remediation.
• If possible where combination remedies have been used, the
relative contributions of dredging, capping and backfilling, and natural
recovery should be measured through sediment monitoring, and the results of monitoring should be used to adapt and optimize remedies.
• Remediation decision makers should examine the expected net
risk reduction associated with each remedial alternative before selecting
a remedy that will be implemented and link the monitoring program to
the assessment of net risk reduction for the selected remedy.
227
• Pre-remediation baseline monitoring methods and strategies
should be developed to allow statistically valid comparisons with future
monitoring datasets that rely on time-series data. The ultimate goal is to
assemble a consistent long-term dataset for conducting evaluations. During preliminary and final remedy design, monitoring should be initiated
to help to establish a time trend integrating earlier characterization data
as technically appropriate.
• Monitoring of the benthic community, as surrogate species reflecting food-web transfer, can provide valuable information about ecosystem health and integrate short-term and long-term exposures and
multiple life stages. When consistent with the conceptual site model of
exposure and risk development of site-specific approaches is recommended.
• Faster and less expensive monitoring methods that are deployable in the field are needed to better inform dredging operations in real
time and to improve predictive capability.
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6
Dredging at Superfund Megasites:
Improving Future Decision-Making
INTRODUCTION
The preceding chapters discussed sediment management and
dredging at Superfund megasites and included sections on assessing the
effectiveness of dredging for removing contaminated sediment to attain
remedial-action objectives and achieve specified cleanup levels. The assessment included the review of 26 projects from which general conclusions were developed with respect to the appropriate use and limitations
of dredging in meeting risk-based goals. From those conclusions, the
committee developed guidelines with respect to favorable site conditions
under which dredging should be more likely to achieve long-term remedial-action objectives. The committee also offered recommendations for
monitoring to facilitate scientifically based and timely decision- making
to improve dredging effectiveness.
In this final chapter, the committee addresses the charge in the
statement of task to consider “how conclusions about completed and
current operations can inform future remedial decision-making” and to
“develop recommendations that will facilitate scientifically based and
timely decision making for megasites in the future.” Specifically, we seek
to identify how lessons learned from experience may inform future prac-
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tices and management of contaminated sediment at megasites. This includes the expected role of dredging in the future and the issues and factors that need to be addressed to ensure the effective use of dredging as a
component of contaminated sediment remediation. Most of the committee’s earlier recommendations focus on these issues at the site-specific
level, but this chapter focuses on the national level.
MANAGING SEDIMENT MEGASITES IN THE FUTURE
With the establishment of Superfund in 1980, we now have the opportunity for retrospective analysis at dozens of sediment sites to evaluate decision-making, field experience, and remedial effectiveness where
dredging has been selected for sediment cleanup. In the past, a rigorous
evaluation of whether site remediation achieved risk reduction goals and
what factors contributed to or limited the achievement of those goals
was often just not done. Although information is available from various
sites with respect to volume of bulk material removed or sediment concentration achieved, that information does not permit determination of
the degree to which remedial objectives for risk-reduction were
achieved. Thus, it is not easy to determine which approaches resulted in
risk reduction under various site conditions. The difficulty stems partly
from the lack of comprehensive post-dredging monitoring data and from
the fact that followup assessments typically do not quantify uncertainty
in both risk measurements and predictions.
In hindsight, it is clear that there are limitations to dredging effectiveness. With this historical perspective comes the opportunity to learn
and improve how we think about and implement environmental dredging. Perhaps nothing is more important than to step back and derive
common lessons from experience, as was done in Chapter 4. This type of
review needs to be continuous and needs to part of a shared experience
among regulators, practitioners, and the public.
As described in Chapter 2, sediment megasites are among the most
challenging and costly sites on the National Priorities List (NPL).
Megasites are conventionally defined as sites with remedial activities
costing at least $50 million; the U.S. Environmental Protection Agency
(EPA) has defined contaminated sediment megasites as sites for which
the sediment component of remedial activities will cost at least $50 million.
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The charge to this committee focused on megasites, but the dearth of
such sites with completed dredging remedies and with good predredging and post-dredging data has meant that the committee reviewed smaller sites, or individual projects at megasites. The projects
evaluated did not include any of the magnitude (that is, tens of miles of
river stretches and thousands of hectares) and time frames (for example,
decades in the Hudson River [TAMS Consultants, 2000]) that can be anticipated at the largest of the current and future megasites. Megasites
with a broad spatial area and large volume of contaminated sediments
will likely require multiple seasons of dredging.1 The larger scale and
time frames will increase the chemical exposures and residual production related to operations, and make it more difficult to fully characterize
contaminant distribution, sources of contamination, and conditions unfavorable for remedial operations. At these sites, risk based goals may
not be achievable in the foreseeable future due to the long time frame,
complexity of the sites, and the limitations of available technologies. The
committee recognizes that experience with remediation at larger sites
might reveal challenges not faced at the smaller sites and has attempted
to anticipate such challenges to the extent possible in making its recommendations.
Cleanup of contaminated sediment megasites incorporates large
temporal and spatial scales that create two distinct issues: the human
health and ecosystem risk-reduction benefits achieved by isolated remediation in a large-scale watershed are difficult to predict and quantify;
and the large spatial scales and long time lines, coupled with the complexity and heterogeneity of large-scale megasites, suggest that varied
and combined remedial approaches will be appropriate. We can do a
better job of addressing those issues by taking a broader, basin-wide
view in contaminated sediment management and by embracing more
flexible approaches. Those issues are discussed further below.
The Need for Regional-Scale Perspectives
Because contaminated sediment megasites are influenced by regional-scale phenomena, watershed and airshed contributions to sediFor example, at current levels of operation, it will take more than 25 years to
complete dredging at New Bedford (Dickerson and Brown 2006).
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ment contamination at any site must be viewed in a larger framework to
permit valid predictions about cleanup and risk reduction. Sediment
megasites can span an entire waterbody (such as the Lower Fox River in
Wisconsin) or be located in a watershed amongst other contaminated
sediment sites. Bridges et al (2006) comment that several watersheds in
the United States contain multiple contaminated sediment sites in close
proximity to one another and that effective sediment management will
require a more holistic approach to understanding multiple sources of
contaminants (sediments, outfalls, and non-point sources) and their cumulative impacts in a waterway.
The public has a right to know what benefits will be achieved for
particular investments. For example, if the risk being addressed is associated with the consumption of fish, a valid question is, How much will
contaminants in fish decrease as a result of this action? Some organisms
travel great distances and there is need to understand their movement
and variable exposure to the Superfund sites. At the same time, the longdistance movement of toxics from the site throughout the larger water
body needs to be understood. Finally, there needs to be an understanding of secular changes in basin-wide conditions and how they might relate to cleanup at a specific site. (For example, whether basin-wide concentrations of contaminants are declining because of point source
reduction or whether contaminants are migrating in from other contaminated areas in the wider basin.) These factors contribute heavily in
evaluating site specific data on concentrations in fish species and the
broader water body before and after a cleanup.
As such, a regional approach to modeling and analysis of contaminants at megasites is needed to better understand their effect on resident
and migratory fish and on the flux of contaminants within the wider basin (e.g. Linkov et al. 2002; von Stackelberg et al. 2002). Because of the
difficulty in accurately estimating several of the necessary parameters
and inputs for these types of models (particularly fish exposure to contaminated sediments, differences in movement of various fish populations or life stages, contaminant concentrations in prey, and uptake and
loss kinetics of mobile species occupying areas of high and low exposure), their uncertainty will remain a concern. However, because these
issues are of particular importance at megasites and where multiple
Superfund sites exist in close proximity, the development of these models is essential.
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A related issue is the lack of essential tools for understanding how
reductions in sediment toxicity or biologic exposure will enhance ecosystem response and benefit ecosystem recovery. Much of our understanding of these topics is wholly observational or is derived from ancillary
measures (such as sediment chemistry). In itself, such information provides little capability for predicting community or organism response
after remediation. Understanding the ecosystem dynamics that affect
recovery entails larger regional-scale phenomena, such as larval recruitment, food-web interactions, and fate and transport processes.2
The Use of Adaptive Management at Megasites
Given the difficulty in predicting dredging effectiveness, and the
limited number of available alternative technologies, what changes can
be made to improve the remedy selection and implementation process to
ensure more effective and cost-effective remedies?
A major challenge to decision-makers is the uncertainty about
whether—and how well—a remedy will work at a site. Experience has
shown the wisdom of well-designed pilot field tests and experimentation
prior to committing to a specific final cleanup remedy. Pilot testing, including monitoring of appropriate environmental variables, for example
as part of the feasibility study, is used to test the performance of a technology or approach, understand the factors affecting its performance,
and to provide information on how, if necessary, the remedy should be
adapted to achieve desired goals. In this way, the information generated
in the pilot tests and monitoring becomes a key component of the remedy selection, design, and implementation process.
The use of a structured process of selecting a management action,
monitoring the effects of the action, and applying those lessons to optimize a management action is generally referred to as adaptive management (e.g., NRC 2003, 2004, 2005; Bridges et al. 2006; Linkov et al 2006a,
b). As described in NRC 2004, “There is no prototype for its implementation, and no ‘cookbook’-type set of steps or building blocks that will immediately constitute an adaptive-management program. It is contextspecific, it involves feedback and learning between scientists, managers,
As discussed in Chapter 5, biodynamic approaches that are linked with principles of functional ecology can help to bridge this knowledge gap.
2
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and stakeholders.” NRC (2005) recommends an adaptive-management
approach at Superfund megasites where it is unlikely that final remedies
can be identified and implemented. That report describes adaptive management as a six-step interactive process for defining and implementing
management policies under conditions of high uncertainty regarding
results of remedial actions (see Box 6-1).
Bridges et al. (2006) described the need for greater flexibility in
sediment management processes because “the more strictly linear decision making process characterized as the ‘decide and defend’ approach
to remedial decision making does not contain sufficient flexibility” and is
unable to accommodate or benefit from other approaches such as adaptive management. In the current Superfund process, the ROD is the end
result of a long and often difficult and contentious process of conducting
studies and receiving and responding to input from stakeholders often
with divergent and impassioned views of the type and extent of remediation that is required (see Chapter 2 for greater detail on the remedy
selection process). A ROD often selects a specific remedy and predicts its
ability to achieve cleanup levels and remedial action objectives at the
site. Because the scale of megasites is so large, a variety of unanticipated
conditions can greatly influence the results of a remediation. When
remedies are selected without the benefit of actual, on-the-ground feedback on the effect of the remediation, there is a greater chance that unforeseen conditions and events will hinder progress or limit the effectiveness of the remediation. The ROD process can be reopened to amend
or modify a ROD on the basis of information gathered after implementation begins, however, instituting an adaptive-management process from
the outset recognizes the uncertainty inherent in predicting remedial
results and allows adaptation of the remedy based on site experience to
optimize progress toward attaining remedial goals.
In this process, the primary goals of Superfund, the protection of
human health and the environment, remain paramount. As such, adaptive management is not a means to permit or sanction less rigorous
cleanups, or to avoid public input or scrutiny of the decision making
process. The principles of transparency and public notification remain
essential and the adaptive-management process at a site needs to be developed in concert with stakeholders and insights from monitoring and
testing need to be shared with them so that they can contribute to
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BOX 6-1 Six-Step Adaptive-management Process
1. Assessing the problem, including establishing measurable management
objectives, key indicators of those objectives, quantitative or conceptual models
to predict effects of remedial alternatives on the indicators, and forecasts of responses of indicators to remedial actions.
2. Designing a management plan, including comparing and selecting remedial actions and, importantly, selecting indicator values that will trigger a
change in management actions.
3. Implementing the plan, including documenting and agreeing with
stakeholders on those circumstances that might require deviations from the plan.
4. Monitoring for effectiveness and for verifying and updating the conceptual model.
5. Evaluating results obtained from monitoring, including comparing results with forecasts from earlier modeling, seeking to explain why results occurred, and provide recommendations for future action.
6. Adjusting the management plan in response to the monitoring results,
including implementing recommendations, reviewing and updating models,
and developing new forecasts, management objectives, and management actions
as necessary.
Source: Adapted from NRC 2005.
adapting the remedy, if necessary. It is expected that adaptive management could be implemented in the current legislative framework because
CERCLA and the NCP have great flexibility and do not preclude adaptive management (NRC 2005). That implementation would need to be
reviewed by EPA to best fit CERCLA requirements.
There is progress toward implementing adaptive management at
contaminated sediment sites. Recent EPA guidance (EPA 2005) endorses
the general concept, stating
Project managers are encouraged to use an adaptive management
approach, especially at complex sediment sites to provide additional certainty of information to support decisions. In general, this
means testing of hypotheses and conclusions and reevaluating site
assumptions as new information is gathered.
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There are also examples of sites where adaptive-management principles, if not an explicit, rigorous adaptive-management process, have
been applied in remediating contaminated sediment sites (see Box 6-2).
In sum, the desired outcome of this more flexible approach is to allow, indeed to encourage, adaptation to realities on the ground in an effort to achieve remedial goals in as efficient and cost effective manner as
possible. These suggested changes reflect the need to make decisions in
the face of uncertainty while allowing managers and stakeholders to respond to, and take advantage of, unanticipated events and a variety of
possible future outcomes through the design of a flexible, iterative learning process.
BOX 6-2 Examples of the Application of Adaptive-management
Principles in Sediment Remediation
In the Fox River, WI, two demonstration projects were conducted during
the Remedial Investigation/Feasibility Study at Sediment Management Units
56/57 and Deposit N (Foth and Van Dyke, 2000; Montgomery Watson, 2001). The
projects provided useful information on implementability, effectiveness, and
expense of large-scale dredging at the site and were used to inform future decision-making.
In Operable Unit (OU) 1 of the Fox River, the ROD permits flexibility in
achieving cleanup levels and stipulates additional actions (further dredging or
capping) if dredging doesn’t achieve desired results. Following dredging, if sampling shows that the 1 ppm action level has not been achieved, a surfaceweighted average concentration (SWAC) of 0.25 ppm may be used to assess the
effectiveness of PCB removal. If that SWAC of 0.25 ppm has not been achieved
for OU 1, the first option is that additional dredging may be undertaken to ensure that all sediments with PCB concentrations greater than the 1 ppm action
level are removed throughout the particular deposit. A second option is placing
a sand cover on dredged areas to reduce surficial concentrations to achieve a
SWAC of 0.25 ppm for OU 1 (WI DNR/EPA 2002).
Finally, in the case of the Grasse River, several large-scale dredging and
capping projects have revealed site-specific conditions that limited the effectiveness of the remediation (including dredging and capping). This site-specific information can then be used in development of a revised Analysis of Alternatives
Report (Alcoa Inc. 2005).
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The Future of Dredging
While some improvements have been made to dredge design and
operation (for example, precision positioning systems or dredge head or
bucket modifications to reduce resuspension), in many respects dredging
as a technology has not changed dramatically in the last few decades.
What has changed is how dredging is applied. Devices designed and
proven for navigational or maintenance dredging are now pressed into
service for specific and precise contaminant-mass removal or to attain
specific sediment contaminant concentrations in what are often complex
settings and difficult conditions. In addition, it is often difficult to accurately characterize the sites, and define the degree of uncertainty about
the effectiveness of different remedial approaches.
The committee found that most of the sites that it examined exhibited one or more conditions unfavorable for dredging and concluded
that dredging alone is unlikely to be effective in achieving both shortterm and long-term cleanup levels at many sites. However, its effectiveness as a contaminant-mass removal technology will ensure its use at
most sites where mass removal is necessary (such as where navigational,
source reduction, or sediment stability concerns are present). Where unfavorable conditions exist, it is likely to be implemented—in conjunction
with capping, in situ treatment, or monitored natural recovery—as part
of a combined remedy. In the future, dredging will continue to play an
important role in the management of Superfund megasites and should
be viewed as one of several approaches that may be necessary for their
cleanup.
CONCLUSIONS AND RECOMMENDATIONS
The committee envisions that some combination of dredging, capping or covering, and natural recovery will be involved at all megasites.
In situ treatments may also be required at many sites. Thus, all remedial
approaches should be considered in the site evaluation, and the interactions among the various approaches should be well understood. Dredging for mass removal itself may be attractive from the viewpoint of the
public, but it alone does not necessarily produce risk reduction. A better
appreciation of the existing risks before dredging and what is required to
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achieve desired risk reduction, both in the short term and in the long
term, is needed.
The challenge to this committee was twofold: to make pertinent
technical recommendations (contained in Chapters 4 and 5) and to recommend changes in the management of the Superfund program that will
ensure that the technical recommendations are implemented. The committee believes that three kinds of changes are critical to improve decision-making and increase dredging-remedy effectiveness at contaminated sediment megasites.
First, owing to the complexity, large spatial scale, and long time
frame involved, the management of contaminated megasites should embrace a more flexible and adaptive approach to accommodate unexpected conditions and events, new knowledge, technology changes, and
results of field pilot tests.
Second, improved risk assessment should specifically consider the
full range and real-world limitations of remedial alternatives to allow
valid comparisons of technologies and uncertainties.
Third, EPA needs a centralized focal point for coordinated assessment of contaminated sediment megasites for better consistency in site
evaluations, remedy selection, and for increased focus and communication among EPA management and technical staff on what works and
why.
Similar recommendations have been discussed and developed by
other groups3, but it is hoped that in the aggregate the committee’s rec-
3The notion that large, complex sites need a more adaptive approach to remedy selection and implementation was the topic of much discussion at the meetings of the EPA National Advisory Council for Environmental Policy and Technology Superfund Subcommittee (NACEPT 2004). The need for adaptivemanagement approaches at complex contaminated sites has also been discussed
in the academic community for some years (for example, Cannon 2005). Various
National Research Council committees and other independent reviews have advocated similar approaches. For example, the National Research Council advocated the use of flexible phased implementation and adaptive management in
environmental remediation (NRC 2001, 2003, 2005), recommended that the wide
array of risks associated with implementing a remedy be explicitly considered
(NRC 2001, 2005), and recommended that the limitations associated with dredging and the potential for production of residual contaminated sediment be con-
250
ommendations add more specificity than past efforts regarding the effective remediation and management of contaminated sediment. The three
recommendations, which are described in more detail below, will in
some cases require additional resources and, equally challenging in large
organizations, new ways of doing business. The committee cannot stress
enough that because of the potentially huge cost and the complexity of
sediment megasites, the costs and efforts required to change standard
operating procedures are worth the up-front investment that will be required. As noted, many times cleaning up sediment megasites sites may
take decades from investigation to cleanup and cost hundreds of millions
of dollars. The cost of implementing the recommendations in this report
should be viewed in that context. In fact, the committee is concerned that
if its recommendations are not implemented, many hundreds of millions
of dollars of government and private funds will be wasted on ineffective
remedies for contaminated sediment megasites.
1. An adaptive management approach is essential to the selection
and implementation of remedies at contaminated sediment megasites
where there is a high degree of uncertainty about the effectiveness of
dredging.
If there is one fact on which all would agree, it is that the selection
and implementation of remedies at contaminated sediment sites are
complicated. Many large and complex contaminated sediment sites will
take years or even decades to remediate and the technical challenges and
uncertainties of remediating aquatic environments are a major obstacle
to cost-effective cleanup.
Because of site-specific conditions—including hydrodynamic setting, bathymetry, bottom structure, distribution of contaminant concentrations and types, geographic scale, and remediation time frames—the
remediation of contaminated sediment is neither simple nor quick, and
the notion of a straightforward “remedial pipeline” that is typically used
to describe the decision-making process for Superfund sites is likely to be
at best not useful and at worst counterproductive.
sidered (NRC 1997, 2001). Yet, at the time of the present review, little progress
has been made in implementing those recommendations.
251
The typical Superfund remedy-selection approach, in which site
studies in the remedial investigation and feasibility study establish a single path to remediation in the record of decision, is not the best approach
to remedy selection and implementation at these sites owing to the inherent uncertainties in remedy effectiveness. At the largest sites, the time
frames and scales are in many ways unprecedented. Given that remedies
are estimated to take years or decades to implement and even longer to
achieve cleanup goals, there is the potential—indeed almost a certainty—
that there will be a need for changes, whether in response to new knowledge about site conditions, to changes in site conditions from extreme
storms or flooding, or to advances in technology (such as improved
dredge or cap design or in situ treatments). Regulators and others will
need to adapt continually to evolving conditions and environmental responses that cannot be foreseen.
These possibilities reiterate the importance of phased, adaptive approaches for sediment management at megasites. As described previously, adaptive management does not postpone action, but rather supports action in the face of limited scientific knowledge and the
complexities and unpredictable behavior of large ecosystems [NRC,
2004].
2. EPA should compare the net risk reduction associated with the
various remedial alternatives, taking into account the limitations of
each approach in selecting site remedies, such as residuals and resuspension.
One subject of great interest and concern at contaminated sediment
Superfund sites is the risk-based comparison of remedial alternatives to
support selection of a remedy (Bridges et al. 2006; Wenning et al. 2006).
The committee was charged only with evaluating the effectiveness of
dredging and not with comparing the effectiveness of remedial alternatives. However, the committee recognizes that the effectiveness of a
dredging remedy depends on good planning, and good planning includes an evaluation of net risk reduction associated with each remedial
alternative. Therefore, the committee recommends evaluating the net
risk reduction of remedial alternatives to facilitate scientifically based
decision making at megasites.
252
Baseline risk is quantified in the remedial investigation for all NPL
sites, but the feasibility study may or may not include a quantitative estimate of the risks posed by alternative remedies. EPA (2005, p. 2-14) indicates that “although significant attention has been paid to evaluating
baseline risks, traditionally less emphasis has been placed on evaluating
risks from remedial alternatives, in part because these risks may be difficult to quantify.” Even if such quantitative comparative risk assessment
is provided, it might be limited in scope. For example, the feasibility
study for the Upper Hudson River (TAMS Consultants 2000) included a
quantitative comparison of human health and ecologic risk reduction for
the fish-consumption exposure pathway, the pathway associated with
the highest risk estimates. However, the analysis did not quantify shortterm effects on the local community or workers or other effects that
might occur during dredging; it concluded that “there is no reliable
means of quantifying potential short-term impacts from activities such as
sediment resuspension, habitat loss, or other transient effects.”
Each remedial alternative offers its own set of risk-reduction benefits and possibly the creation of new exposure pathways and associated
risks. A confounding issue is that site conditions can change in ways that
help to reduce risk. That would be the case, for example, with deposition
of cleaner material over residual contamination. Site-specific measurements and models need to incorporate an understanding of such site features both spatially and temporally to support valid comparisons of remedial alternatives.
Environmental responses to remediation, including sediment and
biota concentration changes, are complex and difficult to predict. During
remedy selection, the uncertainty around estimates of responses to
remediation should be recognized and quantified to the extent warranted to optimize decision-making. For example, EPA established a
tiered approach to probabilistic risk assessment in the Superfund program, as shown in Figure 6-1 (EPA, 2001). Using that approach, one proceeds from a less expensive point estimate sensitivity analysis to more
expensive and time-consuming quantitative uncertainty-analysis methods. The question is: When are the more advanced methods useful or
necessary? Box 6-3 illustrates a situation in which additional quantitative
analysis might be warranted. It presents an idealized comparison of risk
estimates, including inherent uncertainty, for two remedial alternatives.
253
Tier 3 Advanced PRA
2-D MCA
Tier 2 PRA
1-D MCA
Probabilistic Sensitivity Analysis
Complete RI/FS Process
Increasing Complexity/Resource Requirements
Characterization of Variability and/or Uncertainty
Probabilistic Sensitivity Analysis
(Microexposure Modeling, Bayesian Statistics,
Geostatistics)a
Tier 1 Point Estimate Risk Assessment
Point Estimate Sensitivity Analysis
Problem Formulation/Scoping/Work Planning/Data Collection
= Decision Making Cycle: Evaluation, Deliberation, Data Collection,
Work Planning, Communication
At each tier, a decision may be to exit the tiered process
aExamples of advanced methods for quantifying temporal variability, spatial variability, and
uncertainty.
FIGURE 6-1 EPA’s tiered approach to the use of probabilistic risk assessment
(PRA). DMCA: decision-making cycle analysis. Source: EPA 2001.
Some potential effects will remain difficult to accurately quantify
and compare (for example, the impact of a large dredging project on
quality of life issues such as noise or light pollution) or potential psychological consequences from not implementing a removal remedy (for example, if community members perceive that an unmitigated threat to
human health exists in their environment). Other “implementation risks”
(risks potentially imposed by the implementation of a remediation strategy) such as worker and community health and safety, equipment failures, and accident rates associated with an active remediation are given
254
little consideration in EPA’s feasibility studies at Superfund sites (Wenning et al. 2006). Cura et al. (2004) identify several challenges associated
with comparative risk assessment, given data limitations and the unavoidably subjective nature of quantifying some risks associated with
dredged-material management decisions. However, ignoring those types
of risk in comparisons of remedial options is not the solution and may
have undesirable consequences, particularly when the cost of being
wrong is high (Bridges et al. 2006).
BOX 6-3 Importance of Quantifying Uncertainty in Risk Estimates
In the hypothetical case outlined in the figure below, remedial alternative 1
appears to result in lower risk than alternative 2. If the uncertainty in these risk
estimates is not quantified in some way, a risk manager might proceed with alternative 1. However, the uncertainty in this estimate is sufficiently high that one
cannot be certain about this conclusion, and in fact a higher risk might result
from implementing alternative 1. Such an outcome is obviously not desirable
and even more problematic if alternative 1 is the more costly remedial alternative. With the benefit of the quantitative uncertainty analysis, and depending on
the magnitudes of risk estimates and remedy costs, a risk manager might elect to
gather more data (for example, with a pilot field test) to reduce uncertainty in
the risk estimate for remedial alternative 1 before making a selection. A quantitative uncertainty analysis can reveal significant contributors to uncertainty in risk
estimates, which are the most useful subjects of further study and data collection.
Error bars
reflect
uncertainty
about best
estimate of risk
Best
Estimate
of Risk
Remedial
Alternative
#1
Remedial
Alternative
#2
Hypothetical comparison of risk predicted for two remedial alternatives, including quantification of uncertainty associated with the risk estimates.
255
3. There is a great need for centralized EPA resources, responsibility, and authority at the national level to ensure that necessary improvements are made so that contaminated sediment megasites are
remediated as effectively as possible.
As discussed in Chapter 2 and elsewhere in this report, it became
abundantly clear during the committee’s work that EPA has not devoted
adequate resources and senior management attention to the issue of contaminated sediment, given the scope of the problem and the huge costs
incurred by the federal government, the private sector, and others. If the
recommendations in this report (and the many good reports that have
gone before) are to be successfully implemented, some group in the
Superfund program should be given the resources and responsibility to
make needed changes and should then be held accountable.
EPA is in the best position to gather and evaluate relevant data on a
national level, so it is natural for EPA to lead the effort in monitoring the
progress and sharing experiences on dredging at megasites. Because
every EPA region has on-the-ground experience with dredging at some
megasites, regular review and shared experience can inform decisionmaking and raise the overall level of technical expertise. Whether by a
more robust Contaminated Sediments Technical Advisory Group or
some other mechanism, a consistent set of design and monitoring principles should emerge and grow from such efforts. Such information
should be publicly available. The goal is generating a greater understanding of sound remediation principles and best practices and their
uniform application among sites. The difficulty that the committee had
in obtaining information about Superfund contaminated sediment sites
and the lack of consistent data on those sites point to a need for a much
stronger national program that has the authority and responsibility for
overseeing and evaluating EPA’s Superfund contaminated sediment efforts. The recommendations made here and by many earlier independent
evaluations are unlikely to be implemented by the current patchwork
approach to managing contaminated sediment sites. Resources, authority, and strong leadership are needed to ensure that the recommendations in this and prior reports are implemented in a timely manner.
It is impossible to identify a focal point for contaminated sediment
sites in the current Superfund office organizational structure. Yet those
sites are among the most challenging and expensive sites on the NPL.
256
Years of experience suggests that to garner the needed resources, focus,
and management attention for a problem of this magnitude, it is necessary not only to create a “critical mass” of personnel and expertise but to
clearly identify those responsible and accountable for implementing
needed changes and policies. The committee strongly recommends that
this gap be addressed.
Specific responsibilities include the following:
• Gather data to define the scope of the contaminated sediment
problem.
• Track current and likely future contaminated sediment
megasites that are on the NPL and in other EPA programs.
• Review site studies, remedies, and monitoring approaches at
contaminated sediment megasites to assess whether best practices are
being implemented, including whether regions are complying with national sediment and other program guidance.
• Ensure that adaptive-management approaches are applied at
contaminated sediment megasites where there is substantial uncertainty
about the effectiveness of dredging and other remedial approaches. As
part of this effort, it is critical that EPA staff communicate clearly to local
citizens and other stakeholders objective information about what dredging and other remedial options can and can not accomplish, as well as
inform them about the inherent uncertainties of remedy effectiveness at
sediment sites. For an adaptive-management approach to be successful,
public involvement should occur “early and often.”
• Ensure adequate pre- and post-remediation monitoring at complex contaminated sediment sites.
• Evaluate the effectiveness of sediment remediation in near “real
time” at major sediment cleanup projects to determine whether selected
remedies are achieving their intended goals and to develop lessons
learned.
• Create a centralized, easily accessible, and up-to-date repository
of relevant data and lessons learned regarding sediment remedies, including dredging and other approaches to facilitate information transfer
among regional and headquarters staff working on these sites and the
public.
• Develop and implement a research strategy for evaluating ways
to improve the assessment, monitoring, and cleanup of contaminated
257
sediment sites, including the development and testing of new technologies.
• Serve as a focal point for coordination and communication
among the many EPA programs and federal agencies who are involved
in the cleanup of contaminated sediment sites.
For these functions to be implemented, EPA Headquarters program
staff will need to review and provide input to regional decisions, and
present senior program managers with data and information about remedy effectiveness and the approaches tried at different sites on an ongoing (and real time) basis. This will require the commitment of senior
mangers in EPA Headquarters and the 10 EPA regional offices to work
together.
Some of the specific tasks that will be needed to implement the
above functions are described in more detail below.
EPA should define the scope of the problem.
One of the necessary tasks will be to define the scale of the
megasite-sediment problem. As noted in Chapter 1 of this report, EPA
has attempted to define the extent of contaminated sediment in the
United States since at least the 1970s. The latest report (EPA 2004), based
on the National Sediment Inventory (NSI) database, surveys about 9% of
the water-body segments in the United States and classifies 43% of this
nonrandom sample as having probable associated adverse effects. However, EPA’s efforts have fallen short of the systematic assessment needed
to define the scope of contaminated sediment that may require remedial
action.4 Even at Superfund sites, EPA’s efforts to determine the geographic extent and volume of contaminated sediment appears episodic,
Similarly, a recent EPA Office of the Inspector General report (EPA 2006,
P.19) concluded that “EPA’s 2004 National Sediment Quality Survey report did
not provide a complete assessment of the extent and severity of sediment contamination across the Nation, nor did it fully meet the requirements of the Water
Resources Development Act. . . . As a result, EPA cannot accurately estimate the
volume and risks posed by contaminated sediments on a national scale. Such a
national assessment would better enable EPA to ensure that it devotes resources
to contaminated sediment issues that pose the greatest risks to human health and
the environment.”
4
258
and, as described in Chapter 2, there is no current list of contaminated
sediment sites, nor does the Agency evaluate new NPL sites when they
are listed to develop a “watch list” of those sites that are likely to be future megasites. In this regard, conclusions of one of EPA’s earliest reports on the subject (EPA 1987) still holds true: “Although it is reasonable to say that there is significant in-place contamination in U.S. waters,
it is not possible with the current level of knowledge to quantify the
problem. We do not know and cannot even begin to estimate, for example, the river miles affected or the cubic yards of sediment involved.” Of
course EPA can not and should not wait until it has compiled a definitive
picture of the contaminated sediments problem in the United States to
move forward with site cleanups.
Defining the scope of the sediment problem is important for two
reasons. First, it will help to place the magnitude of the problem in
proper perspective to help in understanding how much of the problem
has already been addressed and how much remains to be done. A concrete goal for EPA should be to have an on-going process of evaluating
newly listed NPL sites, as well as major contaminated sediment sites addressed by other programs, to understand the magnitude and severity of
contaminated sediments. From those evaluations, the agency should
produce a report that describes the number of past, active, and probable
future contaminated sediment Superfund sites and the number of likely
megasites. The report should describe the types of contaminants and the
volumes of contaminated sediment and lay out an estimate of likely future costs of cleanup and long-term monitoring. This kind of information
was not available from EPA, which made it difficult to understand the
scope of the problem.
Second, documenting how much work remains to be done and at
what cost should help senior EPA management and other officials to
identify the most pressing program and research needs. For example, if
many more site remedies remain to be executed or listed on the NPL, it
makes sense to invest in developing new technologies for remediation. If
few remedies remain to be chosen, then it may be that developing monitoring tools is of greater importance.
All the recommendations in this report will take staff time and
money to implement. By clearly defining the scope of the problem, EPA
management will have the information it needs to identify the most im-
259
portant tasks to accomplish the goal of improving the scientific basis of
selecting the most effective remedies for contaminated sediment sites.
EPA should ensure that adequate monitoring strategies are implemented.
As highlighted in Chapters 4 and 5, one difficulty in understanding
the effectiveness of dredging is that statistically valid baseline predredging condition assessments are generally not done. Without adequate pre-dredging and post-dredging monitoring, it is impossible to
make the valid comparisons that are necessary to support definitive
statements about the degree to which remedial objectives have been attained as a result of dredging. Much greater attention should be given to
sufficient monitoring to allow valid statistical comparisons of conditions
before and after dredging. That will require considerable forethought in
sampling design, sampling methods, and analytic techniques. The long
period from site investigation through remedial action to required 5-year
reviews compounds the problem.
All that points to the need for EPA (and hopefully other federal
agencies with a stake in this arena, for example, the Army Corps and the
Navy) to invest in better and more consistent measuring tools to monitor
conditions in the field more reliably and efficiently. The committee recommends greater efforts to develop better methods to measure sediment
stability and transport processes, biogeochemical processes, and pore
water concentrations and fluxes.
EPA should develop and implement a contaminated sediment research and
evaluation strategy.
One of the key elements of an improved sediment-cleanup program is to establish a coherent research and evaluation strategy to fill
critical information gaps. In Chapter 4, the committee reached some specific conclusions regarding factors that contribute to or limit dredging
effectiveness. The EPA research and evaluation strategy should build on
the work of the committee to ensure that experience gained in dredging
in a variety of combinations and situations is translated into useful guidance to EPA regions and communicated to the full panoply of external
stakeholders in a timely and transparent fashion.
260
The objective of the research and evaluation program should be to
answer key questions as to what risk reduction will be achieved by different technical approaches, under what site conditions, and with what
certainty. The agency needs to answer those questions through pilot
studies and data collection efforts that monitor baseline to long-term
conditions and that stress robust sampling and statistical analysis. To
this end, EPA should undertake or commission real-time independent
evaluations of the effectiveness of dredging and other remedies at contaminated sediment sites, especially megasites. The reviews would build
on the committee’s analyses and assess the effectiveness of dredging and
other remedies at all major sediment cleanup projects and seek to understand the factors that contributed to or limited the effectiveness of the
cleanup approach. This kind of study should either be conducted by a
neutral external organization (either academic or non-profit) or, if conducted by EPA, be made subject to external peer review. It should be
clear at the outset that an external organization conducting the review
will have full control of the results, and that the final report will be made
publicly available.
This type of systematic evaluation will require EPA’s Superfund office and Office of Research and Development to work together to fill the
critical gaps in guidance and standard protocols. This effort should also
involve other agencies, such as the U.S. Geological Survey, the Army
Corps of Engineers, the U.S. Navy, and the National Oceanic and Atmospheric Administration, who work on contaminated sediment sites.
To implement a unified research and evaluation strategy successfully,
EPA will need to ensure that appropriate resources are applied and that
the various EPA and other government offices involved are held accountable for timely implementation of the strategy once it has been developed.
While there are only a few general approaches to sediment remediation, there is room for improvement in their performance. Improving
and optimizing remediation systems has long been a cornerstone in environmental engineering and remediation. The refinement, modification,
and development of sediment remediation approaches and technologies
can overcome limitations to remedial performance and improve effectiveness. Therefore, research to improve and develop new remediation
technologies, site-characterization techniques, and monitoring tools is
essential to advance sediment remediation and should be supported.
261
Efforts to understand and promote those practices and operations that
improve remediation effectiveness, and the training of decision makers
and practitioners in those operations, is also critical to advance the field
and improve the performance of remedial operations.
In sum, EPA’s efforts should focus on moving forward with remedies at sites and, at the same time on investing the effort needed to make
sure that each new pilot test or remedy implemented increases our collective knowledge of what works and what does not work and why. Because many of contaminated sediment sites are vast and remediating
them will be expensive, it is worth investing time and resources now to
try to ensure more cost-effective remedies in the future. Such a focus is
needed if the country is to make the best possible use of the billions of
dollars that will be spent on site remediation.
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Appendix A
Statement of Task for the Committee on
An NRC committee will conduct an independent evaluation of
dredging projects that will look at the expected effectiveness of dredging
contaminated sediments at Superfund megasites. The assessment will
consider whether EPA’s estimated risk reduction benefits are likely to be
achieved in the time frame as predicted. Aspects of risk reduction include decreased potential for current and long-term exposure of human
and ecological receptors and decreased potential for environmental dispersion of contaminants. The assessment will also consider the potential
for short-term increases in risks due to resuspension during dredging.
The committee will consider sites where information is available for assessing dredging effectiveness. It will strive to develop recommendations that will facilitate scientifically based and timely decision making
for megasites in the future. In doing so, the committee will consider
whether current monitoring regimens are sufficient to inform assessments of effectiveness and what practices should be implemented in
monitoring strategies. The committee will not recommend particular remedial strategies at specific sites. The committee’s considerations will
include:
• Whether planned sediment cleanup levels have been reached
and maintained after dredging.
264
Appendix A
265
• If the predicted magnitude and timing of risk reduction as a result of dredging are likely to be achieved.
• The key site-specific factors that contribute most to achieving
high dredging effectiveness.
• The short-term and long-term impacts on ecologic communities
as a result of dredging.
• Monitoring strategies in use and proposed for use at dredging
sites and whether these strategies are sufficient to inform assessments of
effectiveness.
• The specific types of assessments useful for measuring effectiveness, in particular, measuring the reduction of risk.
• How conclusions about completed and ongoing dredging operations can inform decisionmaking in the future.
It is expected that sources of information available for this assessment would include megasites for which dredging has been completed;
megasites for which plans have been developed; partially implemented,
and operations are ongoing; and smaller sites that exhibit lessons relevant to megasites.
Appendix B
Biographic Information on the
Committee on Sediment Dredging at
Superfund Megasites
Charles O'Melia (Chair) is the Abel Wolman Professor of Environmental
Engineering and chair of the Department of Geography and Environmental Engineering at Johns Hopkins University, where he has served
on the faculty for over 25 years. Dr. O'Melia's research fields include
aquatic chemistry, environmental colloid chemistry, water and wastewater treatment, modeling of natural surface and subsurface waters, and
the behavior of colloidal particles. He has served on the advisory board
and review committees for the environmental engineering departments
of multiple universities. He has advised professional societies, including
the American Water Works Association and Research Foundation, the
Water Pollution Control Federation, the American Chemical Society, and
the International Water Supply Association. He served the U.S. Environmental Protection Agency as a peer-review panel member and on the
Science Advisory Board. Dr. O'Melia has consulted for a variety of municipal, industrial, and government clients. In addition, he has served on
several National Research Council committees, including chairing the
Steering Committee for the Symposium on Science and Regulation and
the Committee on Watershed Management for New York City. He was
also a member of the National Research Council Water Science and
266
Appendix B
267
Technology Board and Board on Environmental Studies and Toxicology.
Dr. O'Melia earned a PhD in sanitary engineering from the University of
Michigan. In 1989, Dr. O'Melia was elected to the National Academy of
Engineering for significant contributions to the theories of coagulation,
flocculation, and filtration leading to improved water-treatment practices
throughout the world.
G. Allen Burton is a professor of environmental sciences and director of
the Institute for Environmental Quality at Wright State University. He
has served as a NATO senior research fellow in Portugal and a visiting
senior scientist in Italy and New Zealand. He was the Brage Golding Distinguished Professor of Research at Wright State University. Dr. Burton's
research during the last 25 years has focused on developing effective
methods for identifying ecologic effects and stressors in aquatic systems
where sediment and storm-water contamination is a concern. His ecosystem risk assessments have evaluated multiple levels of biologic organization, from microbial to amphibian effects. Dr. Burton serves on numerous national and international scientific committees, review panels,
councils, and editorial boards, and he consults for industry and regulatory agencies. He earned his PhD in environmental science (aquatic toxicology) from the University of Texas at Dallas.
William Clements is a professor at Colorado State University, where he
has served on the faculty since 1989. Dr. Clements’s primary research
interests are in basic aquatic ecology and ecotoxicology. His research has
focused on understanding how benthic macroinvertebrate communities
respond to natural and anthropogenic stressors. More recently, his research projects have included assessments of recovery from fire disturbance, quantifying interactions between natural and anthropogenic
stressors, and measuring abiotic factors that influence contaminant
bioavailability. Dr. Clements has a substantial record of publication on
benthic invertebrates, benthic community interactions, and effects of
stressors. He is the author of several book chapters and a coauthor of the
book Community Ecotoxicology, published in 2002. Dr. Clements earned
his PhD in zoology from Virginia Polytechnic Institute and State University in 1988.
268
Frank C. Curriero in an assistant professor in the Departments of Environmental Health Sciences and Biostatistics at the Bloomberg School of
Public Health, Johns Hopkins University. His research expertise and interests include applications of spatial statistics and geographic information systems for environmental public health. Dr. Curriero’s research has
spanned applications involving environmental epidemiology, disease
mapping, spatial variation in risk and exposure assessment models, and
geostatistical methods. His current methodologic research includes statistical methods for censored spatial data and models for non-Euclidean
isotropic spatial dependence in geostatistics. Dr. Curriero earned his
PhD in statistics from Kansas State University.
Dominic Di Toro is the Edward C. Davis Professor of Civil and Environmental Engineering in the Department of Civil and Environmental
Engineering at the University of Delaware and is a consultant for HydroQual, Inc. Dr. Di Toro has specialized in the development and application of mathematical and statistical models to stream, lake, estuarine,
and coastal water-quality and sediment-quality problems. He has participated as an expert consultant, principal investigator, and project
manager on numerous water- quality studies for industry, research
foundations, and government agencies. Recently, his work has focused
on the development of water-quality and sediment-quality criteria,
sediment-flux models for nutrients and metals, and integrated hydrodynamic, sediment-transport, and water-quality models. Dr. Di Toro received his PhD in civil and geological engineering from Princeton University. In 2005, Dr. Di Toro was elected to the National Academy of
Engineering for leadership in the development and application of
mathematical models for establishing water-quality criteria and making
management decisions.
Norman Francingues retired from the U.S. Army Corps of Engineers in
2002 with over 30 years of federal civil service. He is the recipient of the
Army Engineer Association Bronze Order of the de Fluery Medal and
the Army Meritorious Civilian Service Award from the chief of engineers. He is a senior consultant with OA Systems Corporation. Mr.
Francingues worked for the Army Corps as a senior technical adviser
and for other national and international agencies on the environmental
engineering aspects of navigation and hazardous-waste projects. He was
Appendix B
269
technical lead for the development of innovative dredging technologies
for the Army Corps of Engineers Dredging Operations and Environmental Research (DOER) program. He advises on contaminated dredged
material for the International Navigation Association (PIANC), headquartered in Brussels, Belgium. His research involves innovative dredging technologies, fluidized-sediment evaluations, confined placement of
contaminated dredged material, and treatment of contaminated sediments and soils. Mr. Francingues earned an MS in environmental engineering from Mississippi State University.
Richard Luthy is the Silas H. Palmer Professor and chair of the Department of Civil and Environmental Engineering at Stanford University.
His research interests include environmental engineering and water
quality, particularly phase partitioning and the treatment and fate of hydrophobic organic compounds. His research emphasizes interdisciplinary approaches to the behavior and availability of organic contaminants
and the application of these approaches to bioavailability and environmental-quality criteria and sediment restoration. He chaired the National
Research Council's Water Science and Technology Board and its Committee on Bioavailability of Contaminants in Soils and Sediments. He is a
past president of the Association of Environmental Engineering Professors. He is a registered professional engineer and a Diplomate of the
American Academy of Environmental Engineers. He received his PhD in
environmental engineering from the University of California, Berkeley.
Dr. Luthy was elected a member of the National Academy of Engineering in 1999 for leadership in the treatment of industrial wastewaters,
contaminated soils, and aquifers.
Perry L. McCarty is the Silas H. Palmer Professor Emeritus in the Department of Civil and Environmental Engineering at Stanford University. He directed the Western Region Hazardous Substance Research
Center from 1989 to 2002. Dr. McCarty specializes in environmental engineering with emphasis on biologic processes for water-quality control
and the control of hazardous substances in treatment systems and
groundwater. His research interests over the last 45 years have been in
biologic processes for the control of environmental contaminants. His
early research was on anaerobic treatment processes, biologic processes
for nitrogen removal, and biologic degradation of hazardous chemicals.
270
His current interests are in aerobic and anaerobic biologic processes for
control of chlorinated solvents, advanced wastewater-treatment processes, and movement, fate, and control of groundwater contaminants.
Dr. McCarty earned his ScD in sanitary engineering from the Massachusetts Institute of Technology. He was elected to the National Academy of
Engineering in 1977 for contributions to the environmental engineering
profession through education, research, and service to government and
industry.
Nancy Musgrove is the president of Management of Environmental Resources, Inc. Ms. Musgrove is experienced as an aquatic ecologist; working with both regulators and the regulated community, she has expertise
in assessment of risks to aquatic communities, water-quality and sediment- quality investigations, and design of environmental monitoring
programs and laboratory and field studies. She has been involved in
numerous regional and national sediment investigation and cleanup projects and the peer review of decisions made at contaminated-sediment
sites. Ms. Musgrove has substantial experience with the regulatory
framework and technical protocols governing environmentalmanagement decisions throughout the United States and Canada. She
earned an M.S. in fisheries from the University of Washington.
Katherine N. Probst is a senior fellow at Resources for the Future. Over
the last 25 years, she has conducted numerous analyses of environmental
programs, focusing mainly on improving the implementation of Superfund and other hazardous-waste management programs. She was the
lead author of the study Superfund's Future: What Will it Cost?, requested
by Congress, on the estimated cost of the Superfund program to the U.S.
Environmental Protection Agency (EPA). Her most recent study, Success
for Superfund, includes recommendations for specific information that
EPA should make available to the public on all Superfund sites in a site
“report card.” Ms. Probst also has investigated issues related to the use
of institutional controls at contaminated sites, long-term stewardship,
and the cleanup of sites in the nuclear-weapons complex. She was a
member of EPA's Superfund National Advisory Council for Environmental Policy and Technology Subcommittee and of the EPA Science
Advisory Board committee that reviewed analyses of the benefits of the
Appendix B
271
Superfund program. Ms. Probst received an MA in city and regional
planning from Harvard University.
Danny Reible joined the faculty of the University of Texas at Austin College of Engineering in 2004; he holds the Bettie Margaret Smith Chair of
Environmental Health Engineering. He is also director of the Hazardous
Substance Research Center/South and Southwest, a consortium of Louisiana State University, Rice University, Texas A&M University, the
Georgia Institute of Technology, and the University of Texas at Austin.
Dr. Reible leads both fundamental and applied efforts in the assessment
and management of risks associated with hazardous substances, especially as they apply to contaminated sediments. Dr. Reible has led the
development of in situ sediment capping, and he has evaluated the applicability of capping technology to a wide array of contaminants and
settings, including polycyclic aromatic hydrocarbons from fuels, manufactured-gas plants, and creosote- manufacturing facilities; polychlorinated biphenyls; and metals. He has consulted for both industry and
regulatory groups on the applicability and design of capping for remediation at a variety of specific sites. His research has also focused on the
natural attenuation of contaminants as a result of various processes in
the environment. He received his PhD in chemical engineering from the
California Institute of Technology. Dr. Reible was elected a member of
the National Academy of Engineering in 2005 for the development of
widely used methods of managing contaminated sediments.
Louis J. Thibodeaux is the Jesse Coates Professor at the Louisiana State
University College of Engineering. Dr. Thibodeaux’s experience and expertise are in chemical-transport processes at and across the natural media (air, water, soil, and sediments) interfaces. Specific applications have
included chemical movement associated with landfill disposal, treatment, and storage of aqueous waste. He has conducted environmental
research projects on chemical spills in rivers, volatiles from wastewater,
nutrient cycling/modeling in lakes, and hazardous substances in contaminated bed sediment in natural aquatic systems. His current research
efforts address three key aspects of the remediation chemodynamics of
bed-sediment contamination: the natural recovery processes of in situ
bed-sediment in the aquatic environment of rivers, lakes, and estuaries;
the processes occurring with the surface soils formed from extracted (ex
272
situ) dredged material; and the chemodynamics associated with mud
clouds produced during dredging. Dr. Thibodeaux has served the National Research Council as cochair of the Steering Committee for the National Symposium on Strategies and Technologies for Cleaning up Contaminated Sediments in the Nation's Harbors and Waterways, the
Committee on Risk-Based Criteria for Non-RCRA Hazardous Waste, the
Committee on Contaminated Marine Sediments, and the Committee on
Remedial Action Priorities for Hazardous Waste Sites. Dr. Thibodeaux
earned his PhD in chemical engineering from Louisiana State University.
Donna J. Vorhees is a principal scientist with the Science Collaborative,
where she provides human health risk-assessment consulting services
for a variety of municipal, federal, and industrial clients. She is also an
instructor at the Boston University School of Public Health, where she
teaches a course in risk- assessment methods. She has extensive experience in addressing environmental questions arising from multipathway
human exposure to chemicals that have been released to indoor and outdoor environments at federal and state hazardous-waste sites. Her research interests include development of probabilistic human-exposure
models; field surveys to collect data needed to support risk assessment,
such as samples of biota consumed by humans and interviews with anglers regarding fish-consumption practices; identification of research
priorities for improving dredged-material management; and preparation
of environmental-health educational materials. Dr. Vorhees conducted
probabilistic analyses of multipathway exposure to PCBs in residences
near the New Bedford Harbor, MA, Superfund site, to PCBs and pesticides that accumulate in fish from an offshore dredged-material disposal
site, and to PCBs, dioxins, and furans that accumulate in agricultural
products from the floodplain of a contaminated river. She is an active
member of the Society for Risk Analysis and the International Society of
Exposure Analysis. Dr. Vorhees earned her master’s degree and doctorate in environmental health from the Harvard School of Public Health.
John R. Wolfe is a senior manager at Limno-Tech, Inc. He has expertise
in fate and transport modeling of contaminants and environmental economics, and he manages projects in contaminated sediment, wastewater
treatment and discharge permitting, combined sewer-overflow control,
and groundwater protection for a variety of municipal, state, federal, and
Appendix B
273
industrial clients. Dr. Wolfe holds an MSE from the University of Michigan, where has an adjunct teaching appointment in the Department of
Civil and Environmental Engineering. He also holds a PhD in economics
from the University of Pennsylvania and was associate professor of economics at Michigan State University. He is a licensed professional engineer and a member of the American Academy of Environmental Engineers.
Appendix C
Summary of Remedial Action Objectives,
Cleanup Levels (Numerical Remedial
Goals), and Their Achievement at
Sediment-Dredging Sites
Note: For additional details on sites, see Table 3-1. For abbreviations, see page
294.
Site: Bayou Bonfouca, LA
Stated Remedial Action Objectives (Related to Sediment Removal): Reduce or
eliminate the potential for ingestion of carcinogens in groundwater, surface soils,
and shellfish. Reduce or eliminate the direct contact threat posed by bayou sediments and onsite surficial creosote waste deposits (EPA 1987).
Stated Cleanup Levels: “Contaminated sediments will be excavated either to a
depth of about 6 in. into the upper cohesive layer or until PAH contamination is
less than 1,300 ppm” (EPA 1987).
Dates of Remediation: 1994-1995.
Were Cleanup Levels Achieved? No chemical confirmation samples immediately after remedy. Later sampling1 met cleanup levels.
Were Remedial Action Objectives Achieved? Partially confirmed.2
1997 sampling by Louisiana (CH2M Hill 2001); 2003 sampling by EPA (EPA
2003a); and 2006 sampling by EPA after Hurricane Katrina (CH2M Hill 2006).
1
274
Appendix C
275
Comments and Lessons Learned: Advances in dredging technology highlighted
ability to dredge sediment accurately. Importance and difficulty of characterizing
contaminant sediment deposits accurately.3 Importance of backfilling. Difficulty
of accessing data (see Chapter 4). Lack of planned post-dredging monitoring.
Less stringent cleanup level (PAHs at 1,300 ppm).
Site: Lavaca Bay, TX
Stated Remedial Action Objectives (Related to Sediment Removal): Not a
CERCLA remedy. Goals of this treatability study were as follows (Alcoa 2000):
develop information to support the technical and economic evaluation of potential remedial actions; evaluate the effectiveness of dredging equipment on removal of mercury impacted sediment in the study area; evaluate potential impacts of dredging on mercury mobilization and residual sediment
concentrations; and, understand the impact that dredging mercury contaminated
sediment may have on mercury levels in Bay biota.
Stated Cleanup Levels: None given.
Comments: Pilot study.
Dates of Remediation: 1998.
Were Cleanup Levels Achieved? Not applicable―no cleanup levels indicated.
Were Remedial Action Objectives Achieved? Pilot-study goals apparently
achieved; not expected to achieve long-term risk reduction.
Comments and Lessons Learned: (1) No significant change in average surficial
sediment contaminant concentrations after dredging. (2) Advantages of pilot
study for describing results of large-scale dredging at this site. (3) Evaluation of
residuals after each dredging pass provided useful information on generation
and concentrations.
Site: Black River, OH
CERCLA remedy. The goal of the sediment remediation project was to remove
PAH-contaminated sediment to eliminate liver tumors in resident brown bullhead populations (Zarull et al. 1999).
Fish sampling was conducted in 1996 and 1997 and resulted in lifting of fishconsumption advisory in 1998 (CH2M Hill 2001), but there is no indication that shellfish sampling has been conducted.
3
“Detailed design investigations during the summer and fall of 1988 [post-ROD,
pre-dredging] showed the volume of contaminated sediments to be approximately
150,000 cy, an increase of three times that estimated in the ROD. This dramatic volume increase resulted in a cost estimate for the selected remedy rising from approximately $55 million to about $150 million” (EPA 1990b).
2
276
Stated Cleanup Levels: No chemical specific cleanup levels given. “The primary
cleanup target was the removal of sediment in the area of the former USS coke
plant to ‘hard bottom,’ or the underlaying shale bedrock. No quantitative environmental targets or end points were established, although post-dredging sampling was required to test for remaining areas of elevated PAH concentrations”
(Zarull et al. 1999).
Were Cleanup Levels Achieved? No quantitative chemical targets, but apparently met operational targets (mass removal and dredging to bedrock).
Were Remedial Action Objectives Achieved? Short-term risk increased, but
long-term risk-reduction targets met.
Comments and Lessons Learned: Need for monitoring of biota. Dredging effective, although uncertain improvement over natural attenuation. Increase in fish
tumors after implementation.
Site: Outboard Marine Corporation (OMC)―Waukegan Harbor, IL
Stated Remedial Action Objectives (Related to Sediment Removal): None
given.
Stated Cleanup Levels: “Sediments in excess of 50 ppm PCB will be removed
from the harbor by hydraulic dredging” (EPA 1984).
Were Cleanup Levels Achieved? No chemical confirmation samples immediately after remediation. Remedy based on assumption that removal of finegrained “muck” overlying glacial till would achieve cleanup levels. Later sampling4 met cleanup levels; current state unclear.5
Were Remedial Action Objectives Achieved? Remedial action objectives not
defined. Fish-tissue concentration trends inconclusive (see text).
Comments and Lessons Learned: Insufficient pre-dredging and duringdredging data on fish concentrations to make comparison with post-dredging
data. Less stringent remedial-action trigger (sediments greater than 50 mg/kg
removed to depth of clean underlying geologic stratum) than for other PCB
cleanups. No chemical verification samples taken on sediment immediately after
dredging.
4
Apparently, sampling in 1996 (EPA 1999) and in 2003 (ILDPH/ATSDR 2004) indicates PCBs in sediments at less than 50 mg/kg; however, sampling events not mentioned in 5-year reviews or site summary submitted to committee.
5
EPA states “OMC Plant 2 is likely a continual source of PCBs to Waukegan Harbor, thus further harbor sediment sampling and analysis is likely needed to confirm
whether cleanup levels are still being met” (EPA 2002).
Appendix C
277
Site: Commencement Bay―Head of Hylebos, Tacoma, WA
Stated Remedial Action Objectives (Related to Sediment Removal): Achieve
“acceptable sediment quality in a reasonable time frame.” Acceptable sediment
quality is defined as “the absence of acute or chronic adverse effects on biological
resources or significant human health risks.” Reasonable time frame was further
defined to be a period of 10 years to allow for natural recovery (via sedimentation) (EPA 1989c).
Stated Cleanup Levels: SQOs were established in 1989 ROD, and PCB value was
modified in 1997 ESD (EPA 1997b).
Were Cleanup Levels Achieved? Cleanup levels met in all but one area; adjacent
nearshore cap was extended to address this area.
Were Remedial Action Objectives Achieved? Dredging operation recently
completed; no long-term data.
Comments and Lessons Learned: Ability to meet cleanup levels under favorable
conditions. Sufficient sampling and reference cores aided effective site characterization and contributed to success of dredging. Capping of one area above
CULs contributed to success. Site sediment characteristics (“soft black muck”
over native material) permitted overdredging and visual characterization of contaminated vs native material. Cost-plus-fee contract incentivized dredging team
to implement BMPs. Pilot testing6 indicated extent and type of debris and issues
related to dredging and dredge-material handling.
Site: Commencement Bay―Sitcum Waterway, Tacoma, WA
Stated Remedial Action Objectives (Related to Sediment Removal): Achieve
“acceptable sediment quality in a reasonable time frame.” Acceptable sediment
quality is defined as “the absence of acute or chronic adverse effects on biologic
resources or significant human health risks.” Reasonable time frame was further
defined to be a period of 10 years to allow for natural recovery (via sedimentation) (EPA 1989c).
Stated Cleanup Levels: SQOs were established in 1989 ROD, and PCB value
modified in 1997 ESD (EPA 1997b).
Were Cleanup Levels Achieved? Cleanup levels met.7
“The pilot program [260 cy removal] provided information on mechanical dredging, offloading of barges to rail cars, and rail transportation of dredged material to
and into an upland landfill” (Dalton, Olmsted & Fuglevand, Inc. 2006).
7
CULs met in all areas immediately after dredging (redredging was conducted in
one subarea) exception one underpier area where natural recovery was able to
achieve SQOs within allowed period (confirmed in 2003) (EPA 2006a [Commencement Bay–Sitcum Waterway, April 26, 2006]).
6
278
Were Remedial Action Objectives Achieved? Not confirmed through biologic
sampling, but long-term monitoring has shown continued (10 years) compliance
with effects-based and risk-based cleanup levels, which were accepted as surrogates for biologic-effects testing. No further monitoring required by EPA at the
site.
Comments and Lessons Learned: Clearly distinguishable contaminated layer
was valuable in achieving remedial action objectives. Combining cleanup with
port redevelopment provided economies for overdredging, ensuring removal.
Compliance evaluated on an area basis used probabilistic criteria (averages and
upper confidence limits) that accommodated single-chemical, noncontiguous,
low-level exceedances after dredging. Inclusion of natural recovery for low-level
contamination in remedial options allowed reasonable response to undredged
inventory in under-pier areas.
Site: Duwamish Diagonal, Seattle WA
CERCLA remedy (Natural Resource Damage Settlement). Restore and replace
natural resources within the lower Duwamish River and Elliott Bay that have
been injured by releases of hazardous materials through remediation of contaminated sediments in the vicinity of combined sewer overflows and storm
drains, source control, and habitat restoration (U.S.A et al. v the City of Seattle,
Consent Decree No. C90-395WD, December 23, 1991).
Stated Cleanup Levels: No chemical specific cleanup levels given. Dredging
performance criteria based on achieved specific elevation for cap placement
(EcoChem Inc. 2005).
Comments: Project viewed as source-control action to address natural resource
damages at a CSO through a hot-spot cleanup, with remaining contamination
addressed as part of Lower Duwamish Superfund site actions.
Were Cleanup Levels Achieved? Dredging performed for cap placement; performance criteria based on elevation specification. Toxicity data used to define
area to be dredged, although performance based on chemical criteria.
Were Remedial Action Objectives Achieved? Unlikely. Cleanup performed as
interim action before selection of remedy for LDW site.
Comments and Lessons Learned: Lack of adherence to BMPs8 resulted in significant transport of contaminated sediment outside the dredge prism. Postdredging monitoring showed increased concentrations of PCBs and other COCs
8
The sediment remediation project closure report indicated that “the most obvious
problems were over-filling the dredge bucket and spilling material out of the bucket
as it was moved to and from the barge.” Water quality monitoring also indicated exceedances of turbidity-compliance criteria (EcoChem Inc. 2005).
Appendix C
279
in adjacent areas, which required placement of additional thin layer of clean material. Biologic monitoring conducted as part of wider LDW site indicates increases in fish-tissue contaminant concentrations at project site (see text).
Site: Puget Sound Naval Shipyard, Bremerton, WA
Stated Remedial Action Objectives (Related to Sediment Removal): Reduce
risks to subsistence fishers consuming seafood from Sinclair Inlet by reducing
PCB concentrations in biologically active zone of sediment in marine Operable
Unit, controlling shoreline erosion of contaminated fill material, and selectively
removing high concentrations of mercury that were colocated with PCBs (EPA
2000c).
Stated Cleanup Levels: PCBs at 0.023 mg/kg wet weight in fish tissue and at 3.0
mg/kg (OC normalized) in sediment on area-weighted average. Sediment remedial action objectives to be achieved within 10 years. No time frame for recovery
of fish tissue (EPA 2006a [Puget Sound Naval Shipyard, May 15, 2006]).
Comments: Action levels were defined to distinguish which technology would
be implemented. Dredging occurred when sediments had PCBs above 12 mg/kg
(OC normalized) or above 6 mg/kg (OC) when mercury was at over 3 mg/kg.
Enhanced natural recovery (thin-layer placement) was applied where PCBs were
at 6-12 mg/kg OC (EPA 2006a [Puget Sound Naval Shipyard, May 15, 2006]).
Were Cleanup Levels Achieved? No immediate post-dredging sampling. Initial
long-term monitoring shows cleanup levels not met.
Were Remedial Action Objectives Achieved? Long-term monitoring shows
sediment quality has not met interim target that would support achieving goals
in desired 10-year period.
Comments and Lessons Learned: Lack of adherence to BMPs. TSS exceedances.
Fish9 and sediment10 contaminant concentrations did not decrease or increased
after dredging. Importance of recognizing issues with entire dredging process
train beforehand.
Fish (English sole) sampling in 2003 indicates that concentrations after dredging
exceed the cleanup goal. Average PCB concentrations are similar to average preremediation concentrations documented by historical (1991-1997) monitoring. The
average mercury concentration is slightly higher than in previous (1994) sampling
(URS 2006).
10
The post-remediation area-weighted average PCB sediment concentration of 7.413 mg/kg of organic carbon (OC) (90th percentile confidence interval) exceeds the preremediation action area-weighted average value of 7.8 mg/kg of OC calculated from
data collected before remediation (URS 2006).
9
280
Site: Harbor Island―Lockheed Shipyard, Seattle, WA
Stated Remedial Action Objectives (Related to Sediment Removal): “Reduce
concentrations of hazardous substances to levels which will have no adverse
effect on marine organisms by eliminating the exposure pathways associated
with residual concentrations of these contaminants. . . . Restore the marine habitat to its most productive condition to the extent practicable. . . . Minimize or
eliminate the potential for recontamination of the cap from groundwater. . . .
Achieve adequate source control to prevent recontamination” (EPA 2006a [Harbor Island Lockheed Shipyard Sediment OU, May 11, 2006]).
Stated Cleanup Levels: Arsenic at 57 mg/kg dry weight, copper at 390 mg/kg
dry weight, lead at 450 mg/kg dry weight, mercury at 0.41 mg/kg dry weight,
zinc at 410 mg/kg dry weight, PCBs at 12 mg/kg organic carbon normalized,
LPAHs at 370 mg/kg organic carbon normalized (low-molecular-weight polynuclear aromatic hydrocarbons), HPAHs at 960 mg/kg organic carbon normalized
(high-molecular-weight polynuclear aromatic hydrocarbons), and tributyltin at
76 mg/kg organic carbon normalized (EPA 1997c, 2003c, 2006a [Harbor Island
Lockheed Shipyard Sediment OU, May 11, 2006]).
Comments: Cleanup levels were based on Washington State Sediment Management Standards (Apparent Effects Thresholds); TBT cleanup level was developed
on basis of site-specific data for protection of invertebrates. Area background
concentrations of PCBs and mercury were allowed to modify boundary (but not
cleanup level within boundary) and define acceptable levels of recontamination.
Were Cleanup Levels Achieved? Cleanup levels not met for metals, PAHs, and
PCBs in some open-water areas that were to be remediated through dredging
only.11 “Enhanced natural recovery” (placement of 6 in. of sand on the sediment
surface) used in some of these areas. No actions taken in two areas with singlechemical, low-level exceedances. Toe of slope at transition from dredging only to
dredging and capping also did not meet CULs. This noncompliant area addressed through overplacement of cap material.
Were Remedial Action Objectives Achieved? No long-term data yet available
on objective of protection and recovery of benthic community health.
Comments and Lessons Learned: Debris affected schedule and cost. Extensive
sediment characterization during dredging included progress cores to assess
adequacy of dredge cuts. Use of test dredge or pilot dredge would have helped
11
The 2005 5-year review (EPA 2005c) stated that “a total of eight sediment samples
were collected from the post-dredge surface of the channel area. . . . All analytical
results were compared to the SQS [sediment quality standards] chemical criteria to
evaluate compliance. . . . [F]rom eight samples, three samples exceeded the SQS for
PCBs only. Three other samples . . . exceeded the SQS for a combination of COCs
[chemicals of concern].”
Appendix C
281
to characterize debris. Experienced contractors successfully completed sediment
handling with careful site management and successfully reduced contaminant
loss. Implementation of BMPs. Technologies (WINOPS) incorporated to permit
successful dredge placement. Change in dredging contracting strategy from production-based to time-and-materials-based contributed to successful remediation.
Site: Harbor Island―Todd Shipyard, Seattle, WA
Stated Remedial Action Objectives (Related to Sediment Removal): “Reduce
concentrations of hazardous substances to levels which will have no adverse
effect on marine organisms by eliminating the exposure pathways associated
with residual concentrations of these contaminants. . . . Restore the marine habitat to its most productive condition to the extent practicable. . . . Minimize or
eliminate the potential for recontamination of the cap from groundwater. . . .
Achieve adequate source control to prevent recontamination” (EPA 2006a, [Harbor Island Lockheed Shipyard Sediment OU, May 11, 2006]).
Stated Cleanup Levels: Arsenic at 57 mg/kg dry weight, copper at 390 mg/kg
dry weight, lead at 450 mg/kg dry weight, mercury at 0.41 mg/kg dry weight,
zinc at 410 mg/kg dry weight, PCBs at 12 mg/kg organic carbon normalized,
LPAHs at 370 mg/kg organic carbon normalized, HPAHs at 960 mg/kg organic
carbon normalized, and tributyltin at 76 mg/kg organic carbon normalized (EPA
1997c, 2003d, 2006a [Harbor Island Lockheed Shipyard Sediment OU, May 11,
2006]).
Comments: Cleanup levels were based on Washington State Sediment Management Standards (Apparent Effects Thresholds); TBT cleanup level was developed
on basis of site-specific data for protection of invertebrates. Area background
concentrations of PCBs and mercury were allowed to modify boundary (but not
cleanup level within boundary) and define acceptable levels of recontamination.
Were Cleanup Levels Achieved? Mercury and PAH cleanup levels not achieved
at a few locations, but concentrations were below action levels12 and thus acceptable without additional remediation.
Were Remedial Action Objectives Achieved? No long-term data yet available
on objective of protection or /recovery of benthic community health.
Comments and Lessons Learned: Extensive sediment characterization during
dredging included progress cores to assess adequacy of dredge cuts. Experienced
12
EPA stated that “the average (mean) concentration and the upper 95% confidence
level on the mean concentration for all COCs are less than SQS chemical criteria for all
analytes. Based on this statistical evaluation, Todd and EPA have concluded that the
post-dredge surface in all areas of the Site meets cleanup criteria” (EPA 2006a [Harbor
Island Todd Shipyards Sediment Operable Unit, May 12, 2006]).
282
contractors successfully completed sediment handling with careful site management to reduce contaminant loss. Implementation of BMPs throughout process
train. Technologies (WINOPS) incorporated to permit successful dredge placement. Use of dredging contractor as consultant during design phase and use of
environmental performance-based contracting contributed to successful remediation.
Site: Cumberland Bay, NY
Stated Remedial Action Objectives (Related to Sediment Removal): Mitigate
the immediate threat to the environment posed by the PCB-contaminated sludge
bed. Rapidly and significantly reduce human and environmental risks. Prevent
further environmental degradation resulting from this known source of PCB
contamination (NYSDEC 1997).
Stated Cleanup Levels: None given.13
Comments: Entire PCB-contaminated sludge bed to be removed (NYSDEC,
1997).
Were Cleanup Levels Achieved? No cleanup levels established.
Were Remedial Action Objectives Achieved? Not determined; some residual
contamination present.
Comments and Lessons Learned: Hardpan, rocks, and gulleys inaccessible to
hydraulic dredge created unfavorable conditions that required multiple dredge
passes and hand-held diver dredging. High residuals after initial dredging; some
contamination remained at termination of project. Lack of quantitative criteria.
Inadequate sampling and characterization techniques limited initial understanding of full extent of contaminated materials.
Site: Dupont―Christina River, DE
Stated Remedial Action Objectives (Related to Sediment Removal): Prevent
exposure to contaminated sediments (EPA 1993a).
Stated Cleanup Levels:
Original site-specific Cleanup
Revised Site-specific Cleanup
Contaminant
Criteriaa
Criteriab
Zinc
5,600 ppm
3,000 ppm
Lead
1,200 ppm
700 ppm
Cadmium
60 ppm
20 ppm
a
From 1993 ROD (EPA 1993a).
b
Original cleanup values were lowered to eliminate need for extensive long-term monitoring program that was part of
1993 ROD (EPA 2005b).
13
From record of decision (NYSDEC 1997): “Question: What is the target level DEC
hopes to achieve of remaining PCB contamination? Response: The NYSDEC has not
set an action level. The goal of remediation is the removal of the sludge bed in its entirety.”
Appendix C
283
Comments and Lessons Learned: These cleanup criteria were apparently used to
delineate area for remediation. In practice, chemical analyses were not used to
verify removal of contaminated sediments. Sediments were removed to the required minimum depth of 2 ft or until underlying stratum was encountered
(URS 1999).
Were Cleanup Levels Achieved? Not determined; no confirmation samples
taken after dredging and backfilling. Removal targets (elevation) were met.
Were Remedial Action Objectives Achieved? Not determined, no confirmation
or long-term monitoring.
Comments and Lessons Learned: Lack of chemical confirmation sampling and
long-term monitoring is problematic.14 Dredging operation was based on removal, not on concentration. Need for source control and a reference site.
Site: Fox River (OU 1), WI
Stated Remedial Action Objectives (Related to Sediment Removal): Achieve, to
the extent practicable, surface-water quality criteria throughout Lower Fox River
and Green Bay. Protect humans who consume fish from exposure to COCs that
exceed protective levels. Protect ecologic receptors from exposure to COCs above
protective levels. Reduce transport of PCBs from Lower Fox River into Green Bay
and Lake Michigan. Minimize downstream movement of PCBs during implementation of remedy (WI DNR/EPA 2002).
Stated Cleanup Levels: Dredge all sediment with PCBs at over 1 ppm or achieve
a surface-weighted average concentration (SWAC) of 0.25 ppm (WI DNR/EPA
2002) (see Comments for explanation).
Comments: If after dredging is completed for OU 1, sampling shows that the 1ppm remedial action level (RAL) has not been achieved, a SWAC of 0.25 ppm
may be used to assess effectiveness of PCB removal. If that SWAC has not been
achieved, the remedy provides options to reduce risk further. The first option is
additional dredging to ensure that all sediments with PCBs at over 1-ppm RAL
are removed throughout the particular deposit. The second option is to place a
sand cover on dredged areas to reduce surficial concentrations so that a SWAC of
0.25 ppm for OU 1 is achieved (WI DNR/EPA 2002).
Dates of Remediation: 2004-present.
Were Cleanup Levels Achieved? Dredging not complete; some subunits have
not achieved desired cleanup levels.
Were Remedial Action Objectives Achieved? Dredging not yet completed.
14
Second 5-year review for the site indicates that vegetative cover and vegetation
species composition are monitored. However, there is no systematic monitoring of
chemical concentrations in sediment, water, or biota and no evaluation of toxicity end
points.
284
Comments and Lessons Learned: Baseline monitoring, although extensive, is not
sufficient to inform long-term monitoring because it began after dredging had
begun at the site. Thin layer of highly contaminated sediment and residuals have
limited success at reaching 1 ppm. Heterogeneity of deposits creates difficulties
in defining the dredge prism. ROD permits flexibility in achieving cleanup levels
and stipulates additional actions (further dredging or capping) if dredging does
not achieve results.
Site: Fox River (Deposit N), WI
CERCLA remedy. Demonstration-project objectives were environmental dredging to remove contaminated sediment to specifications; protection of the river,
local properties, and residents during sediment removal; safe transport and disposal of sediment; and maintenance of good local relations during the project
(Foth and Van Dyke 2000).
Stated Cleanup Levels: Average residual thickness no more than 3 in. in West
Lobe and no more than 6 in. in East Lobe (Foth and Van Dyke 2000).
Were Cleanup Levels Achieved? Target elevations met.15
Were Remedial Action Objectives Achieved? Pilot-project goals met. Not expected to achieve long-term risk reduction.
Comments and Lessons Learned: Pilot project indicated mass removal can be
achieved. Release of PCBs during dredging.16 Bedrock limited ability to dredge
Project summary stated that “project specifications did not require either total
removal of the sediment or removal to a specific PCB sediment concentration as these
sediments rested on a fractured bedrock surface, preventing a dredge cut into a clean
underlying layer” (Foth and Van Dyke 2000).
16
“During dredging in 1998 (Phase I), the upstream average reported PCB water
column concentration was 3.2 ng/l compared to the average downstream PCB water
column concentration of 11 ng/L. The variation between the upstream PCB water
column concentration and the downstream PCB water column concentration measured during dredging reflects an average increase downstream of 3.5 times the upstream value. Similar water column PCB results were obtained during Phase II and III
in the 1999 dredge season. For the 1999 dredge period, the average upstream PCB
water column concentration was 14 ng/L compared to the average downstream PCB
water column concentration of 24 ng/L. This variation represents an increase of 1.7
times the upstream reported value. It can be concluded from this data that dredging
caused an increase in PCB concentrations downstream of the dredge site” (Malcolm
Pirnie, Inc. and TAMS Consultants, Inc. 2004).
15
Appendix C
285
completely, and residual layer was left. Post-dredging concentrations were similar to that before dredging.17
Site: Fox River (SMU 56/57), WI
CERCLA remedy. Demonstration-project objectives were to evaluate potential
effects of large-scale dredging of PCB-contaminated sediments on the Fox River,
to evaluate efficacy of large-scale dewatering and land disposal of PCBcontaminated sediments, and to evaluate potential costs of large-scale dredging,
dewatering, and land disposal of PCB-contaminated sediments (Montgomery
Watson 2001).
Stated Cleanup Levels: 1999 action: to depths consistent with PCBs at 1 mg/kg of
sediment or less (Montgomery Watson 2001). 2000 action: total PCBs at 1 mg/kg
of sediment or 10 mg/kg with at least 6 in. of clean sand backfill (Fort James Corporation et al. 2001).
Were Cleanup Levels Achieved? Cleanup levels not met in first season of dredging (1999); met in 2000 and then backfilled.
Were Remedial Action Objectives Achieved? Pilot-project goals were met; not
expected to achieve long-term risk reduction.
Comments and Lessons Learned: Residual mass and PCB concentrations resulted in redredging and backfilling. Difficulties experienced in dewatering and
solids handling indicated value of pilot studies and of considering full train of
treatment. Dredging released PCBs to water column despite silt curtain controls.
Using turbidity as an indicator of PCB transport is insufficient.18 No sampling
after sand backfilling, so final surface concentrations are not known.
Site: Ketchikan Pulp Company, Ward Cove, AK
toxicity of surface sediments.
“Using the 1998 data, collected just prior to dredging, the pre-dredge average
PCB sediment concentration in Deposit N was 16 ppm, with a maximum concentration of 160 ppm. The post-dredge average PCB sediment concentration in Deposit N
was 14 ppm, with a maximum of 130 ppm” (Foth and Van Dyke 2000).
18
“The TSS and PCB comparison (downstream minus upstream) illustrates that
TSS is not a reliable indicator of PCB transport during a dredging operation. For example, from September 1 to October 6, a period of negative TSS loading (less at the
downstream than at the upstream site), the PCB loading was positive. Thus, if one is
to monitor PCB transport during a remediation operation, sole reliance on turbidity
or TSS measurements is inadequate” (Steur 2000).
17
286
Enhance recolonization of surface sediments to support a healthy marine benthic
infaunal community with multiple taxonomic groups (EPA 2000b).
Comments: Health of benthic communities is to be assessed through toxicity
testing and benthic community analyses as part of long-term monitoring.
Were Cleanup Levels Achieved? Success of cleanup to be determined by toxicity
testing and benthic community analysis. Concentration-based cleanup levels not
defined.
Were Remedial Action Objectives Achieved? Monitoring is continuing; there is
some initial success in reducing benthic toxicity, which was the desired objective.
Comments and Lessons Learned: Effectiveness of backfilling in reducing toxicity
was demonstrated in comparison with locations without backfilling.19 Site demonstrated use of toxicity assays and benthic community analyses as a useful indicator of ecologic improvement after remediation. Dredging and backfilling conducted in a small area compared to backfilling only and natural recovery areas.
Site: Newport Naval Complex―McCallister Point Landfill, RI
human ingestion of shellfish impacted by sediments with COC concentrations
exceeding cleanup levels. Prevent exposure of aquatic organisms to sediments
with COC concentrations exceeding cleanup levels. Prevent avian-predator ingestion of shellfish impacted by sediments with COC concentrations exceeding
cleanup levels. Minimize migration of sediments with COC concentrations exceeding selected PRGs to offshore areas and previously unaffected areas of Narragansett Bay. Prevent washout of landfill debris into marine environment (EPA
2000d).
Stated Cleanup Levels: Copper, 52.9 ppb in pore water; nickel, 33.7 ppb in pore
water; anthracene, 513 ppb in sediment; chrysene, 1,767 ppb in sediment; fluorene, 203 ppb in sediment; total PCBs, 3,634 ppb in sediment (EPA 2000d).
Were Cleanup Levels Achieved? Cleanup levels confirmed analytically immediate after dredging except when bedrock encountered. Long-term monitoring
19
“The 2004 [first long-term monitoring event] data indicated that conditions in the
three thin-layer capping areas had generally improved, while those in three of the
four natural recovery areas generally had not (the shallow natural recovery area was
the exception)” (EPA 2006a, Ketchikan Pulp Company; April 26, 2006).
Appendix C
287
indicates that sediment cleanup levels have been maintained although pore water exceedances and toxicity20 remain.
Were Remedial Action Objectives Achieved? Although narrowly defined, remedial action objectives based on exposure to sediment above cleanup levels
apparently met. Long-term risk reduction inconclusive.
Comments and Lessons Learned: Cleanup levels were met although verification
sampling was insufficient in some locations (hitting bedrock was considered
meeting values). Pore water exceedances of cleanup levels persist after dredging
although source of contamination is not clear.21 Ability to dredge much of site
from shore was advantageous. Incomplete recolonization by shellfish and submerged aquatic vegetation in near term (less than 5 years). Useful and comprehensive range of pre-monitoring and post-monitoring metrics.
Site: GM Central Foundry, St. Lawrence River, NY
Stated Remedial Action Objectives (Related to Sediment Removal): Remedial
action objectives are not specifically provided. EPA does state: “Hot spots in the
St. Lawrence and Raquette rivers and Turtle creek will be dredged and excavated
to remove PCBs. All PCB contaminated sediments in the hot spots will be removed given the technological limitations associated with dredging” (EPA 1991).
Stated Cleanup Levels: St. Lawrence and Racquette Rivers: PCBs at 1 ppm (EPA
1991).
Were Cleanup Levels Achieved? Cleanup levels not met in St. Lawrence River,
because of residuals, backfilling, and capping required in one area. Cleanup levels met in Racquette River.
Were Remedial Action Objectives Achieved? Dredging alone unable to achieve
cleanup levels, although combination remedy effectively reduced surface concentrations.
Comments and Lessons Learned: Intensive monitoring before, during, and immediately after dredging provided useful indications of dredging effect. Sheetpile walls limited PCB release during dredging. Inability to eliminate residuals
and possible increase in residual concentration due to contaminant retention
within sheet-pile walls. Lack of sediment sampling for contamination since
“Sea urchin pore water toxicity results indicate toxicity at some stations within
Groups 1, 2, 3, and 4 (TetraTech NUS, 2006).” (Groups are geographic areas with sampling stations; groups 1, 2, 3, and 4 are all the areas besides the reference area.)
21
The first long-term monitoring report for the site (Tetra Tech NUS 2006) states
that “groundwater discharge may be responsible for elevated pore water metals concentrations found at some of the near shore locations.” However, the committee notes
that dredging may have influenced redox conditions leading to the dissolution of
metals from residual contaminated materials.
20
288
dredging in 1995 eliminates insight into concentration changes over time. No
apparent trend in fish concentrations after dredging.
Site: Grasse River, NY (Non-Time-Critical Removal Action)
Stated Remedial Action Objectives (Related to Sediment Removal): Project
objectives were: (1) remove the most upstream major PCB source in the Grasse
River; (2) eliminate a potential source of PCB exposure to biota; reduce potential
long-term risks to human health and the environment; (3) provide valuable sitespecific data for use in the Analysis of Alternatives for the study area (ALCOA
1995).
Comments: Non-Time-Critical Removal Action (NTCRA).
Were Cleanup Levels Achieved? None provided. Average sediment concentrations decreased.22
Were Remedial Action Objectives Achieved? Majority of sediment removed,
but corresponding reductions in fish-tissue and water-column concentrations not
observed.23
Comments and Lessons Learned: Debris and bedrock created operational difficulties. Site served as useful pilot study; substantial useful data were collected.
Pilot improved site conceptual model. Much of targeted sediment mass was removed with substantial portion of PCBs in river system.24 Removal released
PCBs to water column as evidenced by increased water concentrations and accumulation by caged fish adjacent to work zone. Turbidity release did not correlate with PCB release. Project demonstrated value of caged-fish studies.
According to BBL (1995): “Area A [the largest area] pre-NTCRA samples contained PCB concentrations ranging from non-detect (MDL varied) to 11,000 mg/kg,
while post-removal PCB samples contained PCB concentrations ranging from 1.1 to
260 mg/kg. The arithmetic (geometric) average PCB concentration detected within
area A (considering all depths) was reduced from 1,109 mg/kg (263 mg/kg) to 75
mg/kg (40 mg/kg), or about 93% (85%). Considering PCBs contained within the
bioavailable zone (that is, top 12 in. of sediment), average arithmetic (geometric) PCB
concentrations were reduced from 518 mg/kg (195 mg/kg) to 75 mg/kg (40 mg/kg) or
approximately 86% (80%). At several locations, PCB concentrations in the residual
sediment (all bioavailable) increased from pre-remediation bioavailable conditions.”
In the smaller area B, the pre-NTCRA arithmetic (geometric) average sediment PCB
concentration of 300 mg/kg (275 mg/kg) was decreased to 108 mg/kg (106 mg/kg)
following dredging, approximately a 64% (62%) decrease (BBL 1995).
23
About 84% of the original volume of sediment was removed, leaving 550 cy of
the 3,500-cy sediment that was targeted. Average thickness of sediment in area was
reduced from about 22 in. to 4 in. (BBL 1995).
24
About 27% of PCB mass in Grasse River study area (BBL 1995).
22
Appendix C
289
Site: Grasse River, NY
CERCLA remedy. Demonstration project objectives were (1) evaluate dredging
as a remedial option to reduce the potential risk that may be posed by future ice
jam related sediment scour events by removing a targeted area of sediments with
elevated PCB concentrations in an area of the river that is known to be subject to
ice jam-related scour; (2) develop site-specific information related to dredging
effectiveness, dredging residuals, dredging production rate, and sediment resuspension that can be used in the development of the revised Analysis of Alternatives Report (Alcoa Inc. 2005).
Stated Cleanup Levels: Not a CERCLA remedy. No chemical specific cleanup
levels given. Remove all soft sediments, to the extent possible, from an approximate 8-acre area of the main channel. To the extent practical, all soft sediments
will be removed to hard bottom leaving a stable dredge face on the adjacent
sediments (Alcoa Inc. 2005).
Were Cleanup Levels Achieved? No site-specific cleanup levels.
Were Remedial Action Objectives Achieved? Stated goals were achieved. Not
expected to achieve long-term risk reduction.
Comments and Lessons Learned: Debris and bedrock limited dredging effectiveness. Limited control over backfilling created higher than expected cap concentrations. No significant change in surface concentrations. Significant increase
in some biota followed dredging.
Site: Lake Jarnson, Sweden
CERCLA remedy. Remediation goal was to substantially reduce transport of
PCBs from lake sediments to lake water and downstream system to reduce PCB
concentrations in biota (Fox River Group 1999).
Stated Cleanup Levels: PCBs at maximum of 0.5 ppm and no more than 25% of
remediated area at over 0.2 ppm (Bremle et al 1998b).
Were Cleanup Levels Achieved? Cleanup levels achieved except in one location.25
25
After dredging was completed, 2.9 kg of PCB was calculated to be left in lake
sediment. Sedimentary pool of PCB was distributed, and only one of 54 defined subareas exceeded 0.5 mg/kg dry weight. At commencement of remediation, this had
been set as highest acceptable level to be left in sediment. Only 20% of sediment areas
ended up with PCB higher than 0.2 mg/kg dry weight, which was better than remediation objective of 25% (Bremle et al. 1998b).
290
Were Remedial Action Objectives Achieved? Results indicate reductions in
transport26 and fish-tissue concentrations.27
Comments and Lessons Learned: Ability to overdredge. Sediment concentration
decline corresponds to water and fish-tissue concentration declines. Monitoring
data seek to differentiate regional declines in background PCB concentrations
from those resulting from remediation. Did not target or dredge near-shore
PCBs, which are later implicated as a continuing source of PCBs for fish.28
Site: Manistique Harbor, MI
PCB concentrations in fish and water in the Manistique River and Harbor to levels that would not present an unacceptable human health or ecologic risk and
would allow elimination of existing fish-consumption advisories. Maintain harbor as a navigable waterway for commercial shipping, fishing boats, and recreational watercraft. In general, restore river and harbor areas for use by deeperdraft vessels. Minimize need for future remedial action in area after completion
of a non-time-critical action. Implement actions that would best contribute to
efficient performance of any future remedial actions in the area. Achieve compliance consistent with federal and state ARARs for site. Comply with risk-based
objectives defined by TERRA, Inc., as part of the risk assessment. Reduce, as
much as practicable, the release of PCBs associated with particles and dissolved
in the water to Lake Michigan (EPA 2006a [Manistique River and Harbor Site,
May 10, 2006]).
Stated Cleanup Levels: Initially, goal of action was to remove sediments with
PCB concentrations greater than 10 ppm. Later, goal was modified to state that
objective was 95% removal of total PCB mass and an average sediment concentration not over 10 ppm throughout sediment column (Weston 2002).
26
“After completed remedial activities in Lake Jarnsjon in spring 1995, PCB concentrations had decreased at the outlet of the lake. In addition, the PCB transport during
high flow in the early spring was lower than the year before” (Bremle et al. 1998b).
27
“Fish from all the locations in 1996 [after dredging] had lower PCB concentration
than in 1991 [before dredging]. The most pronounced decrease was observed in the
remediated lake, where levels in fish were halved. The main reason for the reduced
levels was the remediation” (Bremle and Larsson 1998).
28
“Since some of the contaminated sediment was left in the lake after remediation
and this sediment was mainly located in the most shallow, littoral areas, these sites
constituted a probable source of PCB to zooplankton and fish. PCB still remaining in
littoral sediment was probably the cause for a recorded gradient of PCB in fish from
the lake and downstream the river” (Bremle and Larsson 1998).
Appendix C
291
Were Cleanup Levels Achieved? Average cleanup concentration level was met;
it is unclear whether mass-removal goal was met.29
Were Remedial Action Objectives Achieved? Progress toward remedial action
objectives after deposition event.
Comments and Lessons Learned: Initial cleanup levels not met. Poor initial
characterization of wood debris and bedrock issues resulted in incomplete
dredging and a longer dredging time frame. Dredging caused an initial increase
in surface concentrations and no decrease in fish concentrations. Deposition due
to dam removal and sand placement led to decreased surface PCB concentrations.
Site: Reynolds Metals, St. Lawrence River, NY
human and biota contact with contaminated sediments. Reduce or prevent human ingestion of fish caught from the St. Lawrence River. Reduce short-term
effects on surface water and air expected as a result of remedial activities (EPA
2006c).
Stated Cleanup Levels: PCBs, 1 ppm; PAHs, 10 ppm; TDBF, 1 ppb (EPA 1993b).
Were Cleanup Levels Achieved? Cleanup levels for PCBs were met after capping;30 PAH cleanup levels not met; work continues.31
Were Remedial Action Objectives Achieved? Remedial action objectives not
met. PAH remedial activities are ongoing.
Comments and Lessons Learned: Residuals due to bedrock and dredging over
boulders and cobbles. PCB concentrations used to indicate PAH contamination;
however, there was a lack of concordance between PCB and PAH contamination.
“The range of removal efficiency stretches from 82% to 97% depending upon the
many assumptions made in the calculation of the residual mass of PCBs. The outcome
is highly dependant upon the specific gravity assumed for the in-situ sediments along
with sediment volume estimates. The best case estimate indicates that the objective of
95% removal of PCBs from the AOC may have been met while the worst-case estimate indicates that the objective may not have been met.” (Weston 2002).
30
“Despite extensive dredging of the St. Lawrence River, the cleanup goals of 1
mg/kg PCBs, 10 mg/kg PAHs, and 1 μg/kg TDBFs were not achievable in all areas. As
a result, a 0.75-acre, 15 cell area, containing a range of PCB concentrations from 11.1
mg/kg PCBs to 120.457 mg/kg, was capped with the first layer of a three-layer cap to
achieve the cleanup goal [for PCBs]. The remaining exposed sediments average 0.8
mg/kg PCBs within the remaining 255 cells (21 acres), which is below the cleanup
goal” (EPA 2006c).
31
“The remedial action activities in the remaining cells containing elevated levels of
PAHs above the cleanup goal have not been fully implemented” (EPA 2006c).
29
292
Site: Marathon Battery, Hudson River, Cold Spring, NY
cadmium in sediments to protect aquatic organisms and protect human health.
Reduce the transport of suspended sediments from east and west foundry coves
and the pier area (EPA 1989a).
Stated Cleanup Levels: Dredging to 1 ft.
Comments: According to the Record of Decision (EPA 1989a), “The data compiled for east foundry cove indicate that over 95% of the cadmium contamination
is located in the upper layer (1 foot) of the sediments. Due to the nature of the
dredging process, dredging to a specific action level (for example, 10, 100, or 250
mg/kg of cadmium) would be technically difficult, since these concentrations
vary in the sediments by only a few inches of depth. Therefore, expectations are
that by dredging the upper layer of contaminated sediments, 95% of the cadmium contamination will be removed. Following remediation, it is anticipated
that cadmium concentrations would not exceed 10 mg/kg in most of the dredged
areas. . . . Sediment samples at and beneath the cold spring pier will be collected,
analyzed, and evaluated to ascertain whether this area is a source of cadmium
contamination. If, based upon this analysis, these sediments are determined to be
a source, these sediments will be dredged to a depth of one foot.”
Were Cleanup Levels Achieved? Cleanup levels met (dredging performance
targets met).
Were Remedial Action Objectives Achieved? Exposure measures of remedial
action objectives met.
Comments and Lessons Learned: Discrete contaminated layer. Low amounts of
debris. Ability to overdredge.32 Site has useful post-dredging verification sampling and long-term monitoring data on sediments and biota that indicate beneficial effect of remedial activity.
Site: New Bedford Harbor, MA
Stated Remedial Action Objectives (Related to Sediment Removal): Pilotproject objectives were to significantly reduce PCB migration from hot-spot area
sediment, which acts as a PCB source to the water column and to the remainder
of the sediments in the harbor; to significantly reduce the amount of remaining
PCB contamination that would need to be remediated to achieve overall harbor
cleanup; to protect public health by preventing direct contact with hot-spot
32
“EFC [East Foundry Cove] and EFP [East Foundry Pond] bottom sediments consist of silts and clays with some sand. These sediments have a very low bearing capacity which extend to nearly 80 feet in depth. The river bottom sediments consist of silts
and clays with varying amounts of sand” (EPA 2006a [Marathon Battery Superfund
Site, May 10, 2006]).
Appendix C
293
sediments; and to protect marine life by preventing direct contact with hot-spot
area sediments.
Stated Cleanup Levels: Short-term hot-spot goal, 4,000 ppm total PCBs (EPA
1990a); long-term hot-spot goal, 10 ppm total PCBs after additional remediation
occurs (EPA 1998b).
Comments: Owing to limited scope of hot-spot dredging action, EPA did not
expect to achieve standards or levels of control associated with final cleanup levels (such as FDA PCB tolerance for fish tissue and water quality criterion). However, the action was expected to comply with some ARARs, including compliance with RCRA facility regulations, Executive Order 11988 regarding protection
of flood plains to the extent practicable, Executive Order 11990 regarding protection of wetlands, and federal and state air standards during dredging and treatment of contaminated sediments (EPA 1990a).
Dates of Remediation: Hot-spot removal, 1994-1995.
Were Cleanup Levels Achieved? Operations appear to have achieved interim
total-PCB cleanup level of 4,000 ppm (USACE 1995); removal was completed
with minimal net transport of PCBs.
Were Remedial Action Objectives Achieved? Hot-spot removal was not intended to meet long-term risk-reduction goals. Unclear whether hot-spot removal objectives were met.
Comments and Lessons Learned: Usefulness of pilot studies. Value of EPA’s
process modifications on the basis of increased PCB and hydrogen sulfide concentrations in air during sediment handling. Indication of increased exposure to
terrestrial plants and no increased exposure to aquatic biota during dredging.
Challenge of relating contamination and dredging in a large harbor to human
exposure and effects.
Site: United Heckathorn, Richmond, CA
Stated Remedial Action Objectives (Related to Sediment Removal): The ROD
(EPA 1996b) states the clean up goal was based on a surface-water quality criterion for protection of human health from consumption of fish and bioaccumulation of DDT and dieldrin. DDT concentrations exceeded dieldrin concentrations
by a factor of 10-100, so sediment remediation goals for both contaminants were
based on DDT concentrations.
Stated Cleanup Levels: Water column, DDT at 0.59 ng/L and dieldrin at 0.14
ng/L. Sediment, DDT at 590 μg/kg (dry weight) (EPA 1996b).
Comments: Values are based on achieving a 10-6 lifetime excess cancer risk level.
As described in the ROD (EPA 1996b), this value is lower than the chronic marine aquatic-life criterion of DDT at 1 ng/L or dieldrin at 1.9 ng/L. Human health
criteria were judged likely to be achieved if average sediment DDT concentration
was below 0.59 mg/kg (dry weight), which is lower than the 1 mg/kg that would
probably meet marine chronic water quality criteria (EPA 1996b).
294
Were Cleanup Levels Achieved? Cleanup levels apparently met immediately
after dredging. Recontamination after dredging.
Were Remedial Action Objectives Achieved? Little improvement in channel;
remedial action objectives not met.
Comments and Lessons Learned: Dredging was not effective in decreasing
sediment or water concentrations, and biota concentrations did not decline to
clean levels. Side slopes, piers, ship traffic, debris, and an outfall may have led to
increased residual contamination. Difficulty in reaching agreement between parties. Conceptual site model was not sufficient to discern effect of dredging and
likelihood of recontamination. Usefulness of deployed mussel studies to indicate
effect of dredging.33
ABBREVIATIONS: PAH, polycyclic aromatic hydrocarbon(s); CERCLA, Comprehensive Environmental Response, Compensation, and Liability Act; PCB,
polychlorinated biphenyl; SQO, sediment quality objective; ROD, record of decision; ESD, explanation of significant differences; CUL, cleanup level; BMP, best
management practice; LDW, Lower Duwamish Waterway; COC, contaminants
of concern; TSS, total suspended solids; TBT, tributyl tin; PRG, preliminary
remediation goal; ARAR, applicable or relevant and appropriate requirements;
TDBF, total dibenzofurans; RCRA, Resource Conservation and Recovery Act;
DDT, dic.
33
“Year 1 biomonitoring showed that pesticide concentrations in the tissues of
mussels exposed at the site were higher than those observed before remediation. Year
2 samples, collected in February 1999, showed tissue levels that were much reduced
from Year 1 but still exceeded pre-remediation levels of DDT at Lauritzen Channel,
Santa Fe Channel, and Richmond Inner Harbor Channel” (EPA 2006a [United
Heckathorn Superfund Site, Richmond, CA, May 16, 2006]).

Sediment Dredging at Superfund Megasites: Assessing the - CLU-IN

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